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    INTERNATIONAL PROGRAMME ON CHEMICAL SAFETY








    ENVIRONMENTAL HEALTH CRITERIA 165





    INORGANIC LEAD













    This report contains the collective views of an international group of
    experts and does not necessarily represent the decisions or the stated
    policy of the United Nations Environment Programme, the International
    Labour Organisation, or the World Health Organization.

    First draft prepared at the National Institute of Health Sciences,
    Tokyo, Japan, and the Institute of Terrestrial Ecology, Monk's Wood,
    United Kingdom


    Published under the joint sponsorship of the United Nations
    Environment Programme, the International Labour Organisation, and the
    World Health Organization


    World Health Organization
    Geneva, 1995

         The International Programme on Chemical Safety (IPCS) is a joint
    venture of the United Nations Environment Programme, the International
    Labour Organisation, and the World Health Organization. The main
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    the effects of chemicals on human health and the quality of the
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    development of manpower in the field of toxicology. Other activities
    carried out by the IPCS include the development of know-how for coping
    with chemical accidents, coordination of laboratory testing and
    epidemiological studies, and promotion of research on the mechanisms
    of the biological action of chemicals.

    WHO Library Cataloguing in Publication Data

    Inorganic lead.

    (Environmental health criteria ; 165)

    1.Lead - adverse effects  2.Environmental exposure
    3.Guidelines  I.Series

    ISBN 92 4 157165 9                 (NLM Classification: QV 292)
    ISSN 0250-863X

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    CONTENTS

    ENVIROMENTAL HEALTH CRITERIA FOR INORGANIC LEAD

    PREAMBLE

    PREFACE

    1. SUMMARY

         1.1. Identity, physical and chemical properties, and analytical
               methods
         1.2. Sources of human exposure
         1.3. Environmental transport, distribution and transformation
         1.4. Environmental levels and human exposure
         1.5. Kinetics and metabolism in laboratory animals and humans
         1.6. Effects on laboratory animals and  in vitro systems
         1.7. Effects on humans
         1.8. Evaluation of human health risks

    2. IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, AND ANALYTICAL
         METHODS

         2.1. Physical and chemical properties of lead and its compounds
         2.2. Analytical procedures
               2.2.1. Sampling procedures
                       2.2.1.1    Sampling of environmental media
                       2.2.1.2    Sampling of biological materials
               2.2.2. Analytical methods for lead
                       2.2.2.1    Analysis of lead in environmental
                                  samples
                       2.2.2.2    Analysis of lead in biological materials
                       2.2.2.3    Analytical procedures for biomarkers of
                                  lead exposure and effect
         2.3. Conversion factors

    3. SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE

         3.1. Natural occurrence
               3.1.1. Rocks and soils
               3.1.2. Sediments
               3.1.3. Water
               3.1.4. Air
               3.1.5. Plants
               3.1.6. Environmental contamination from natural sources
         3.2. Anthropogenic sources
               3.2.1. Lead mining
               3.2.2. Smelting and refining
               3.2.3. Environmental pollution from production of lead
         3.3. Consumption and uses of lead and its compounds
         3.4. Sources of environmental exposure

    4. ENVIRONMENTAL TRANSPORT, DISTRIBUTION AND TRANSFORMATION

         4.1. Transport and distribution between media
               4.1.1. Atmospheric deposition
               4.1.2. Transport to water and soil
               4.1.3. Transport to biota
                       4.1.3.1    Aquatic organisms
                       4.1.3.2    Terrestrial organisms
         4.2. Environmental transformation
               4.2.1. Abiotic transformation
               4.2.2. Biotransformation

    5. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE

         5.1. Inhalation route of exposure
               5.1.1. Ambient air
                       5.1.1.1    Emissions from motor vehicles
                       5.1.1.2    Stationary sources
               5.1.2. Indoor air
               5.1.3. Air in the working environment
               5.1.4. Smoking of tobacco
         5.2. Exposure by ingestion
               5.2.1. Water
               5.2.2. Food and alcoholic beverages
                       5.2.2.1    Food
                       5.2.2.2    Total intake from food
                       5.2.2.3    Alcoholic beverages
               5.2.3. Dust and surface soils
                       5.2.3.1    Dust
                       5.2.3.2    Soil
                       5.2.3.3    Migration of lead from food containers
         5.3. Miscellaneous exposure
               5.3.1. Cosmetics and medicines
         5.4. General population exposure
         5.5. Blood lead concentrations of various populations
               5.5.1. Adult populations
               5.5.2. Children
               5.5.3. Remote populations
         5.6. Occupational exposure

    6. KINETICS AND METABOLISM IN LABORATORY ANIMALS AND HUMANS

         6.1. Absorption
               6.1.1. Absorption after inhalation
                       6.1.1.1    Animal studies
                       6.1.1.2    Human studies
               6.1.2. Absorption of lead from the gastrointestinal tract
                       6.1.2.1    Animal studies
                       6.1.2.2    Human studies
                       6.1.2.3    Nutritional status and lead absorption
                                  via gastrointestinal tract

               6.1.3. Dermal absorption
                       6.1.3.1    Human dermal absorption
               6.1.4. The relationship of external lead exposure to blood
                       lead concentration
                       6.1.4.1    Ambient air
                       6.1.4.2    Food
                       6.1.4.3    Drinking-water
                       6.1.4.4    Soil and dust
                       6.1.4.5    Total lead intake
         6.2. Distribution
               6.2.1. Animal studies
               6.2.2. Human studies
               6.2.3. Transplacental transfer
         6.3. Elimination and excretion
         6.4. Biological indices of lead exposure and body burden
               6.4.1. Blood lead
               6.4.2. Tooth lead
               6.4.3. Bone lead
               6.4.4. Lead in urine
               6.4.5. Lead in hair

    7. EFFECTS ON LABORATORY ANIMALS AND  IN VITRO TEST SYSTEMS

         7.1. Biochemical effects
               7.1.1. Haem synthesis and haematopoiesis
         7.2. Nervous system effects
               7.2.1. Higher order behavioural toxicity
               7.2.2. Mechanisms of lead-induced behavioural toxicity
                       7.2.2.1    Conclusions
               7.2.3. Sensory organ toxicity
         7.3. Renal system
         7.4. Cardiovascular system
         7.5. Reproductive system
         7.6. Effects on bone
         7.7. Immunological effects
         7.8. Mutagenicity
         7.9. Carcinogenicity

    8. EFFECTS ON HUMANS

         8.1. Biochemical effects of lead
               8.1.1. Haem synthesis
                       8.1.1.1    Protoporphyrin levels
                       8.1.1.2    Coproporphyrin levels
                       8.1.1.3    delta-Aminolaevulinic acid levels in
                                  urine and blood
                       8.1.1.4    Aminolaevulinic acid dehydratase levels
                       8.1.1.5    delta-Aminolaevulinic acid synthase
                       8.1.1.6    Other effects of decreased haem
                                  synthesis

               8.1.2. Vitamin D
               8.1.3. Dihydrobiopterin reductase
               8.1.4. Nicotinamide adenine dinucleotide synthetase
               8.1.5. Nutritionally affected groups
         8.2. Haematopoietic system
               8.2.1. Anaemia
               8.2.2. Pyrimidine-5'-nucleotidase activity
               8.2.3. Erythropoietin production
         8.3. Nervous system
               8.3.1. Historical perspective
               8.3.2. Neurotoxic effects in adults
                       8.3.2.1    Central nervous system
                       8.3.2.2    Peripheral nervous system
                       8.3.2.3    Autonomic nervous system
               8.3.3. Neurotoxic effects in children
                       8.3.3.1    Historical perspective
               8.3.4. Population-based cross-sectional studies on
                       children
                       8.3.4.1    Tooth lead studies
                       8.3.4.2    Blood lead studies
                       8.3.4.3    Follow-up studies
                       8.3.4.4    Conclusions and limitations of
                                  cross-sectional studies
               8.3.5. Prospective epidemiological studies on children
                       8.3.5.1    Common elements
                       8.3.5.2    Study descriptions
                       8.3.5.3    Summary of differences between studies
                       8.3.5.4    Results of studies
                       8.3.5.5    Questions prospective studies have not
                                  answered
                       8.3.5.6    Attempting a consensus
               8.3.6. Task group overview and interpretation of
                       prospective studies on children
                       8.3.6.1    Rationale
                       8.3.6.2    The prospective studies
                       8.3.6.3    A quantitative assessment of the
                                  cross-sectional studies
                       8.3.6.4    Task group overview of cross-sectional
                                  studies
                       8.3.6.5    An interpretation of the overview of
                                  prospective and cross-sectional studies
               8.3.7. Hearing impairment in children
         8.4. Renal system
               8.4.1. Clinical studies
               8.4.2. Epidemiological studies
                       8.4.2.1    Occupational cohorts
                       8.4.2.2    General population
                       8.4.2.3    Cohort mortality studies
         8.5. Cardiovascular system
               8.5.1. Blood pressure

                       8.5.1.1    Studies on occupationally exposed
                                  cohorts
                       8.5.1.2    Studies in the general population
               8.5.2. Other cardiovascular effects
                       8.5.2.1    Occupational studies
                       8.5.2.2    Studies in the general population
               8.5.3. Summary
         8.6. Gastrointestinal effects
               8.6.1. Occupational exposure
               8.6.2. Exposure of children
         8.7. Liver
               8.7.1. Occupational exposure
               8.7.2. Exposure of children
         8.8. Reproduction
               8.8.1. Female populations
               8.8.2. Male populations
               8.8.3. Hormonal responses
               8.8.4. Postnatal growth and stature
         8.9. Effects on chromosomes
         8.10. Carcinogenicity
               8.10.1. Occupational exposure and renal cancer
               8.10.2. Conclusion
         8.11. Effects on thyroid function
               8.11.1. Occupational groups
               8.11.2. Effects in children
         8.12. Immune system
               8.12.1. Occupational exposure
               8.12.2. Children
         8.13. Effects on bone
         8.14. Biomarkers for lead effects

    9. EVALUATION OF HUMAN HEALTH RISKS

         9.1. Exposure assessment
               9.1.1. General population exposure
               9.1.2. Occupational exposures
         9.2. Critical issues related to exposure evaluation
               9.2.1. Sampling and analytical concerns
               9.2.2. Data presentation
         9.3. Relationship between exposure and dose
         9.4. Surrogate measures of dose
               9.4.1. Blood
               9.4.2. Urine
               9.4.3. Bone
               9.4.4. Tooth
               9.4.5. Hair
         9.5. Biochemical effects of lead
               9.5.1. Haem synthesis
                       9.5.1.1    Urinary coproporphyrin
                       9.5.1.2    Urinary aminolaevulinic acid in children

                       9.5.1.3    Urinary aminolaevulinic acid in adults
                       9.5.1.4    delta-Aminolaevulinic acid dehydratase
               9.5.2. Vitamin D metabolism
               9.5.3. Dihydrobiopterin reductase
               9.5.4. Haemopoietic system
                       9.5.4.1    Anaemia in adults
                       9.5.4.2    Anaemia in children
                       9.5.4.3    Erythrocyte pyrimidine-5'-nucleotidase
         9.6. Nervous system
               9.6.1. Adults
                       9.6.1.1    Central nervous system
                       9.6.1.2    Peripheral nervous system
                       9.6.1.3    Autonomic nervous system
               9.6.2. Children
                       9.6.2.1    Type of effect
                       9.6.2.2    Magnitude
                       9.6.2.3    Reversibility/persistence
                       9.6.2.4    Age-specific sensitivity
                       9.6.2.5    Interactions/subgroups
               9.6.3. Animal studies
         9.7. Renal system
         9.8. Liver
         9.9. Reproduction
               9.9.1. Female
               9.9.2. Male
         9.10. Blood pressure
         9.11. Carcinogenicity
         9.12. Immune system

    10. RECOMMENDATIONS FOR PROTECTION OF HUMAN HEALTH

         10.1. Public health measures
         10.2. Public health programmes
         10.3. Screening, monitoring and assessment procedures

    11. FURTHER RESEARCH

    12. PREVIOUS EVALUATIONS BY INTERNATIONAL BODIES

    REFERENCES

    RESUME

    RESUMEN
    

    NOTE TO READERS OF THE CRITERIA MONOGRAPHS

         Every effort has been made to present information in the criteria
    monographs as accurately as possible without unduly delaying their
    publication. In the interest of all users of the Environmental Health
    Criteria monographs, readers are requested to communicate any errors
    that may have occurred to the Director of the International Programme
    on Chemical Safety, World Health Organization, Geneva, Switzerland, in
    order that they may be included in corrigenda.



                               *   *   *



         A detailed data profile and a legal file can be obtained from the
    International Register of Potentially Toxic Chemicals, Case postale
    356, 1219 Châtelaine, Geneva, Switzerland (Telephone No. 9799111).



                               *   *   *



         This publication was made possible by grant number 5 U01
    ES02617-15 from the National Institute of Environmental Health
    Sciences, National Institutes of Health, USA, and by financial support
    from the European Commission.

    Environmental Health Criteria

    PREAMBLE

    Objectives

         In 1973 the WHO Environmental Health Criteria Programme was
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         The first Environmental Health Criteria (EHC) monograph, on
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         of the chemical
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    *    Environmental transport, distribution and transformation
    *    Environmental levels and human exposure

    *    Kinetics and metabolism in laboratory animals and humans
    *    Effects on laboratory mammals and  in vitro test systems
    *    Effects on humans
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    *    Conclusions and recommendations for protection of human health
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         JMPR

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    Procedures

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    FIGURE 1

         It is accepted that the following criteria should initiate the
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    WHO TASK GROUP ON ENVIRONMENTAL HEALTH CRITERIA FOR INORGANIC LEAD

     Members

    Professor S. Araki, Department of Public Health, Faculty of Medicine,
       University of Tokyo, Japan

    Dr P. Baghurst, Division of Human Nutrition, Commonwealth Scientific
       Industrial Research Organization, Adelaide, Australia

    Dr D. Bellinger, Neuroepidemiology Unit, Gardner House, Children's
       Hospital, Boston, Massachusetts, USA

    Dr I. Calder, Occupational and Environmental Health, South Australian
       Health Commission, Adelaide, South Australia, Australia

    Dr D.A. Cory-Slechta, Department of Environmental Medicine, 
       University of Rochester School of Medicine and Dentistry, 
       Rochester, New York, USA

    Dr K. Dietrich, Department of Environmental Health, Division of
       Biostatistics and Epidemiology, University of Cincinnati College
       of Medicine, Cincinnati, Ohio, USA

    Dr R.A. Goyer, Chapel Hill, North Carolina, USA  (Chairman)

    Dr M.R. Moore, University of Glasgow, Department of Medicine and
       Therapeutics, Western Infirmary, Glasgow, Scotland

    Dr C. Nam Ong, Department of Community, Occupational and Family
       Medicine, National University of Singapore, National University
       Hospital, Singapore

    Dr S.J. Pocock, Department of Epidemiology and Population Sciences,
       Medical Statistics Unit, University of London, London, England

    Dr M.B. Rabinowitz, Marine Biological Laboratory, Woods Hole,
       Massachusetts, USA

    Dr M. Smith, Thomas Coram Research Unit, London, England

    Dr G. Winneke, Medical Institute for Environmental Health,
       Heinrich-Heine University, Düsseldorf, Germany  (Vice-Chairman)

     Observers

    Dr C. Boreiko, Environmental Health, International Lead Zinc  Research
       Organization (ILZRO) Inc., Research Triangle Park, North
       Carolina, USA

    Dr N.H. Clark, Lead Industry Environment and Health Forum, Melbourne,
       Victoria, Australia

    Dr J.M. Davis, Environmental Criteria and Assessment Office, US
       Environmental Protection Agency, Research Triangle Park, North
       Carolina, USA

    Professor G. Duggin, Toxicology Unit, Royal Prince Alfred Hospital,
       Camperdown, Australia

    Dr G.R. Neville, Queensland Health Department, Brisbane, Australia

     Secretariat

    Dr G.C. Becking, International Programme on Chemical Safety, 
       Interregional Research Unit, World Health Organization, Research
       Triangle Park, North Carolina, USA  (Secretary)

    Dr K.R. Mahaffey, National Institute of Environmental Health 
       Sciences, Research Triangle Park, North Carolina, USAa

    Dr A.E. Robinson, Toronto, Ontario, Canada  (Rapporteur)

                 

    a  Present address: US Environmental Protection Agency, Environmental
       Criteria and Assessment Office, Cincinnati, Ohio, USA

    ENVIRONMENTAL HEALTH CRITERIA FOR INORGANIC LEAD

         A WHO Task Group on Environmental Health Criteria for Inorganic
    Lead met in Brisbane, Australia, from 1 to 6 February 1993. The
    meeting was sponsored by a consortium of Australian Commonwealth and
    State Governments through a national Steering Committee chaired by
    Dr Keith Bentley, Director, Health and Environmental Policy,
    Department of Human Services and Health, Canberra. The meeting was
    hosted and organized by the Queensland Department of Health,
    Dr G.R. Neville being responsible for the arrangements. Dr G. Murphy,
    Director of Public Health, Queensland, welcomed the participants on
    behalf of the Organizers, and Dr T. Adams, Chief Commonwealth Medical
    Advisor and Dr G. Johns, Parliamentary Secretary to Federal Minister
    for Health, Housing and Community Services, welcomed the participants
    on behalf of the Commonwealth Government. Dr Johns stressed the
    importance attached to this IPCS meeting by the Commonwealth and State
    Governments of Australia. Dr G.C. Becking, IPCS, welcomed the
    participants on behalf of Dr M. Mercier, Director of the IPCS and the
    three cooperating organizations (UNEP/ILO/WHO). The Task Group
    reviewed and revised the draft criteria monograph, and made an
    evaluation of the risks to human health from exposure to inorganic
    lead.

         The Task Group draft was prepared by Dr A.E. Robinson, Toronto,
    Canada, using texts made available by Dr K.R. Mahaffeya (National
    Institute of Environmental Health Sciences, Research Triangle Park,
    North Carolina, USA) and Dr E. Silbergeld (University of Maryland
    School of Medicine, Baltimore, Maryland, USA), and the comments
    received from the IPCS contact points for environmental health
    criteria monographs. The draft was revised extensively by the Task
    Group taking into account the comments from the IPCS contact points.

         Dr G.C. Becking (IPCS Central Unit, Interregional Research Unit)
    and Dr P.G. Jenkins (IPCS Central Unit, Geneva) were responsible for
    the overall scientific content and technical editing, respectively, of
    this monograph.

         The efforts of all who helped in the preparation and finalization
    of this publication are gratefully acknowledged.

                 

    a  Present address: US Environmental Protection Agency, Environmental
       Criteria and Assessment Office, Cincinnati, Ohio, USA

    ABBREVIATIONS

    AAS     atomic absorption spectrometry
    AES     atomic emission spectroscopy
    ALA     delta-aminolaevulinic acid
    ALAD    delta-aminolaevulinic acid dehydratase
    ASV     anodic stripping voltametry
    EDTA    ethylenediaminetetraacetic acid
    FEP     free erythrocyte porphyrin
    GFAAS   graphite furnace atomic absorption spectrometry
    ICP     inductively coupled plasma
    IDMS    isotope dilution mass spectrometry
    MPb     mobilization yield of lead
    MSW     municipal solid waste
    PbB     blood lead
    PbT     tooth lead
    TML     tetramethyllead
    XRFS    X-ray fluorescence spectroscopy
    ZPP     zinc protoporphyrin

    PREFACE

         Although many countries have initiated programmes to lower the
    level of lead in the environment, human exposure to lead remains of
    concern to public health officials worldwide. For over 20 years the
    World Health Organization (WHO) and the International Programme on
    Chemical Safety (IPCS) have been concerned about the health and
    environmental effects of the levels of inorganic lead in the
    environment. The evaluation of human health risks arising from
    food-borne lead has been carried out by WHO on four occasions since
    1972. In addition, health-based guidance values for lead in water, air
    and the workplace have been developed by various Task Groups convened
    by WHO. Environmental Health Criteria 3: Lead, published in 1977,
    examined the effects of lead on human health and Environmental Health
    Criteria 85: Lead - Environmental Aspects was published in 1989.

         Since the publication of Environmental Health Criteria 3: Lead, a
    large body of knowledge has accumulated concerning the effects of lead
    on humans at low levels of exposure. Studies have emphasized the
    effects of inorganic lead on infants and children, a high-risk
    population. This monograph on inorganic lead reflects this research
    emphasis; a major part of the monograph deals with the neurotoxic
    effects of lead with emphasis on neurobehavioural development in
    children. Less detail is presented on the health effects of the higher
    levels of inorganic lead found in some workplaces, although such
    exposures are still considered to pose a risk to humans in many
    regions of the world.

         This monograph deals only with the human health effects of
    inorganic lead. No attempt has been made to evaluate the human health
    effects of organo-lead compounds, although it was recognized that such
    compounds when added to petrol (gasoline) are a major source of
    inorganic lead in the environment. In view of the toxicity of many
    organo-lead derivatives and the possible methylation of inorganic lead
    in the environment, the IPCS plans to evaluate the risk to humans from
    exposure to organo-lead compounds in a separate monograph.

         As with all IPCS criteria monographs, no attempt has been made to
    prepare an exhaustive bibliography of the extremely large amount of
    lead-related literature published since 1977. Rather, an effort has
    been made to review critically the studies on humans and experimental
    animals that are essential for the evaluation of risks to human health
    from exposure to all sources of inorganic lead.

    1.  SUMMARY

         This monograph focuses on the risks to human health associated
    with exposure to lead and inorganic lead compounds. Emphasis has been
    given to data which have become available since the publication of
    Environmental Health Criteria 3: Lead (IPCS, 1977). The environmental
    effects of lead are discussed in Environmental Health Criteria 85:
    Lead - Environmental Aspects (IPCS, 1989).

    1.1  Identity, physical and chemical properties, and analytical
         methods

         Lead is a soft, silvery grey metal, melting at 327.5°C. It is
    highly resistant to corrosion, but is soluble in nitric and hot
    sulfuric acids. The usual valence state in inorganic lead compounds is
    +2. Solubilities in water vary, lead sulfide and lead oxides being
    poorly soluble and the nitrate, chlorate and chloride salts are
    reasonably soluble in cold water. Lead also forms salts with such
    organic acids as lactic and acetic acids, and stable organic compounds
    such as tetraethyllead and tetramethyllead.

         The most commonly used methods for the analysis of low
    concentrations of lead in biological and environmental materials are
    flame, graphite furnace and inductively coupled plasma atomic
    absorption spectroscopy and anode stripping voltametry. Depending on
    sample pretreatment, extraction techniques and analytical
    instrumentation, detection limits of 0.12 µmoles lead/litre blood
    (2.49 µg/dl) can be achieved. However, reliable results are obtained
    only when specific procedures are followed to minimize the risk of
    contamination during sample collection, storage, processing and
    analysis.

    1.2  Sources of human exposure

         The level of lead in the earth's crust is about 20 mg/kg. Lead in
    the environment may derive from either natural or anthropogenic
    sources. Natural sources of atmospheric lead include geological
    weathering and volcanic emissions and have been estimated at
    19 000 tonnes/year, compared to an estimate of 126 000 tonnes/year
    emitted to the air from the mining, smelting and consumption of over 3
    million tonnes of lead per year.

         Atmospheric lead concentrations of 50 pg/m3 have been found in
    remote areas. Background levels of lead in soil range between 10 and
    70 mg/kg and a mean level near roadways of 138 mg/kg has been
    reported. Present levels of lead in water rarely exceed a few
    micrograms/litre; the natural concentration of lead in surface water
    has been estimated to be 0.02 µg/litre.

         Lead and its compounds may enter the environment at any point
    during mining, smelting, processing, use, recycling or disposal. Major
    uses are in batteries, cables, pigments, petrol (gasoline) additives,
    solder and steel products. Lead and lead compounds are also used in
    solder applied to water distribution pipes and to seams of cans used
    to store foods, in some traditional remedies, in bottle closures for
    alcoholic beverages and in ceramic glazes and crystal tableware. In
    countries where leaded petrol is still used, the major air emission is
    from mobile and stationary sources of petrol combustion (urban
    centres). Areas in the vicinity of lead mines and smelters are subject
    to high levels of air emissions.

         Airborne lead can be deposited on soil and water, thus reaching
    humans through the food chain and in drinking-water. Atmospheric lead
    is also a major source of lead in household dust.

    1.3  Environmental transport, distribution and transformation

         The transport and distribution of lead from fixed, mobile and
    natural sources are primarily via air. Most lead emissions are
    deposited near the source, although some particulate matter (< 2 µm
    in diameter) is transported over long distances and results in the
    contamination of remote sites such as arctic glaciers. Airborne lead
    can contribute to human exposures by the contamination of food, water
    and dust, as well as through direct inhalation. The removal of
    airborne lead is influenced by atmospheric conditions and particulate
    size. Large amounts of lead may be discharged to soil and water.
    However, such material tends to remain localized because of the poor
    solubility of lead compounds in water.

         Lead deposited in water, whether from air or through run-off from
    soils, partitions rapidly between sediment and aqueous phase,
    depending upon pH, salt content, and the presence of organic chelating
    agents. Above pH 5.4, hard water may contain about 30 µg lead/litre
    and soft water about 500 µg lead/litre. Very little lead deposited on
    soil is transported to surface or ground water except through erosion
    or geochemical weathering; it is normally quite tightly bound
    (chelated) to organic matter.

         Airborne lead can be transferred to biota directly or through
    uptake from soil. Animals can be exposed to lead directly through
    grazing and soil ingestion or by inhalation. There is little
    biomagnification of inorganic lead through the food chain.

    1.4  Environmental levels and human exposure

         In the general non-smoking adult population, the major exposure
    pathway is from food and water. Airborne lead may contribute
    significantly to exposure, depending upon such factors as use of

    tobacco, occupation, proximity to motorways, lead smelters, etc., and
    leisure activities (e.g., arts and crafts, firearm target practice).
    Food, air, water and dust/soil are the major potential exposure
    pathways for infants and young children. For infants up to 4 or 5
    months of age, air, milk, formulae and water are the significant
    sources of lead exposure.

         Levels of lead found in air, food, water and soil/dust vary
    widely throughout the world and depend upon the degree of industrial
    development, urbanization and lifestyle factors. Ambient air levels
    over 10 µg/m3 have been reported in urban areas near a smelter,
    whereas lead levels below 0.2 µg/m3 have been found in cities where
    leaded petrol is no longer used. Lead intake from air can, therefore,
    vary from less than 4 µg/day to more than 200 µg/day.

         Levels of lead in drinking-water sampled at the source are
    usually below 5 µg/litre. However, water taken from taps (faucets) in
    homes where lead is present in the plumbing can contain levels in
    excess of 100 µg/litre, particularly after the water has been standing
    in the pipes for some hours.

         The level of dietary exposure to lead depends upon many lifestyle
    factors, including foodstuffs consumed, processing technology, use of
    lead solder, lead levels in water, and use of lead-glazed ceramics.

         For infants and children, lead in dust and soil often constitutes
    a major exposure pathway. Lead levels in dust depend upon such factors
    as the age and condition of housing, the use of lead-based paints,
    lead in petrol and urban density. The intake of lead will be
    influenced by the age and behavioural characteristics of the child and
    bioavailability of lead in the source material.

         Inhalation is the dominant pathway for lead exposure of workers
    in industries producing, refining, using or disposing of lead and lead
    compounds. During an 8-h shift, workers can absorb as much as 400 µg
    lead, in addition to the 20-30 µg/day absorbed from food, water and
    ambient air; significant intake may occur from ingestion of large
    inhaled particulate material.

    1.5  Kinetics and metabolism in laboratory animals and humans

         Lead is absorbed in humans and animals following inhalation or
    ingestion; percutaneous absorption is minimal in humans. Depending
    upon chemical speciation, particle size, and solubility in body
    fluids, up to 50% of the inhaled lead compound may be absorbed. Some
    inhaled particulate matter (larger than 7 µm) is swallowed following
    mucociliary clearance from the respiratory tract. In experimental
    animals and humans, absorption of lead from the gastrointestinal tract
    is influenced by the physico-chemical nature of the ingested material,
    nutritional status, and type of diet consumed. In adult humans
    approximately 10% of the dietary lead is absorbed; the proportion is

    higher under fasting conditions. However, in infants and young
    children as much as 50% of dietary lead is absorbed, although
    absorption rates for lead from dusts/soils and paint chips can be
    lower depending upon the bioavailability. Diets that are deficient in
    calcium, phosphate, selenium or zinc may result in increased lead
    absorption. Iron and vitamin D also affect absorption of lead.

         Blood lead (PbB) levels are used as a measure of body burden and
    absorbed (internal) doses of lead. The relationship between blood lead
    and the concentration of lead in exposure sources is curvilinear.

         Once it has been absorbed, lead is not distributed homogeneously
    throughout the body. There is rapid uptake into blood and soft tissue,
    followed by a slower redistribution to bone. Bone accumulates lead
    over much of the human life span and may serve as an endogenous source
    of lead. The half-life for lead in blood and other soft tissues is
    about 28-36 days, but it is much longer in the various bone
    compartments. The percentage retention of lead in body stores is
    higher in children than adults. Transfer of lead to the human fetus
    occurs readily throughout gestation.

         Blood lead is the most commonly used measure of lead exposure.
    However, techniques are now available for measuring lead in teeth and
    bone, although the kinetics are not fully understood.

    1.6  Effects on laboratory animals and in vitro systems

         In all species of experimental animals studied, including
    non-human primates, lead has been shown to cause adverse effects in
    several organs and organ systems, including the haematopoietic,
    nervous, renal, cardiovascular, reproductive and immune systems. Lead
    also affects bone and has been shown to be carcinogenic in rats and
    mice.

         Despite kinetic differences between experimental animal species
    and humans, these studies provide strong biological support and
    plausibility for the findings in humans. Impaired learning/memory
    abilities have been reported in rats with PbB levels of
    0.72-0.96 µmoles/litre (15-20 µg/dl) and in non-human primates at PbB
    levels not exceeding 0.72 µmoles/litre (15 µg/dl). In addition, visual
    and auditory impairments have been reported in experimental animal
    studies.

         Renal toxicity in rats appears to occur at a PbB level in excess
    of 2.88 µmol/litre (60 µg/dl), a value similar to that reported to
    initiate renal effects in humans. Cardiovascular effects have been
    seen in rats after chronic low-level exposures resulting in PbB levels
    of 0.24-1.92 µmol/litre (5-40 µg/dl). Tumours have been shown to occur
    at dose levels below the maximum tolerated dose of 200 mg lead (as
    lead acetate) per litre of drinking-water. This is the maximum dose
    level not associated with other morphological or functional changes.

    1.7  Effects on humans

         In humans, lead can result in a wide range of biological effects
    depending upon the level and duration of exposure. Effects at the
    subcellular level, as well as effects on the overall functioning of
    the body, have been noted and range from inhibition of enzymes to the
    production of marked morphological changes and death. Such changes
    occur over a broad range of doses, the developing human generally
    being more sensitive than the adult.

         Lead has been shown to have effects on many biochemical
    processes; in particular, effects on haem synthesis have been studied
    extensively in both adults and children. Increased levels of serum
    erythrocyte protoporphyrin and increased urinary excretion of
    coproporphyrin and delta-aminolaevulinic acid are observed when PbB
    concentrations are elevated. Inhibition of the enzymes
    delta-aminolaevulinic acid dehydratase and dihydrobiopterin reductase
    are observed at lower levels.

         The effects of lead on the haemopoietic system result in
    decreased haemoglobin synthesis, and anaemia has been observed in
    children at PbB concentrations above 1.92 µmol/litre (40 µg/dl).

         For neurological, metabolic and behavioural reasons, children are
    more vulnerable to the effects of lead than adults. Both prospective
    and cross-sectional epidemiological studies have been conducted to
    assess the extent to which environmental lead exposure affects
    CNS-based psychological functions. Lead has been shown to be
    associated with impaired neurobehavioural functioning in children.

         Impairment of psychological and neurobehavioural functions has
    been found after long-term lead exposure of workers.
    Electrophysiological parameters have been shown to be useful
    indicators of subclinical lead effects in the CNS.

         Peripheral neuropathy has long been known to be caused by
    long-term high-level lead exposure at the workplace. Slowing of nerve
    conduction velocity has been found at lower levels. These effects have
    often been found to be reversible after cessation of exposure,
    depending on the age and duration of exposure.

         The effect of lead on the heart is indirect and occurs via the
    autonomic nervous system; it has no direct effect on the myocardium.
    The collective evidence from population studies in adults indicates
    very weak associations between PbB concentration and systolic or
    diastolic blood pressure. Given the difficulties of allowing for
    relevant confounding factors, a causal relationship cannot be
    established from these studies. There is no evidence to suggest that
    any association of PbB concentration with blood pressure is of major
    health importance.

         Lead is known to cause proximal renal tubular damage,
    characterized by generalized aminoaciduria, hypophosphataemia with
    relative hyperphosphaturia and glycosuria accompanied by nuclear
    inclusion bodies, mitochondrial changes and cytomegaly of the proximal
    tubular epithelial cells. Tubular effects are noted after relatively
    short-term exposures and are generally reversible, whereas sclerotic
    changes and interstitial fibrosis, resulting in decreased kidney
    function and possible renal failure, require chronic exposure to high
    lead levels. Increased risk from nephropathy was noted in workers with
    a PbB level of over 3.0 µmol/litre (about 60 µg/dl). Renal effects
    have recently been seen among the general population when more
    sensitive indicators of function were measured.

         The reproductive effects of lead in the male are limited to sperm
    morphology and count. In the female, some adverse pregnancy outcomes
    have been attributed to lead.

         Lead does not appear to have deleterious effects on skin, muscle
    or the immune system. Except in the case of the rat, lead does not
    appear to be related to the development of tumours.

    1.8  Evaluation of human health risks

         Lead adversely affects several organs and organ systems, with
    subcellular changes and neurodevelopmental effects appearing to be the
    most sensitive. An association between PbB level and hypertension
    (blood pressure) has been reported. Lead produces a cascade of effects
    on the haem body pool and affects haem synthesis. However, some of
    these effects are not considered adverse. Calcium homoeostasis is
    affected, thus interfering with other cellular processes.

    a)   The most substantial evidence from cross-sectional and
         prospective studies of populations with PbB levels generally
         below 1.2 µmol/litre (25 µg/dl) relates to decrements in
         intelligence quotient (IQ). It is important to note that such
         observational studies cannot provide definitive evidence of a
         causal relationship with lead exposure. However, the size of the
         apparent IQ effect, as assessed at 4 years and above, is a
         deficit between 0 and 5 points (on a scale with a standard
         deviation of 15) for each 0.48 µmol/litre (10 µg/dl) increment in
         PbB level, with a likely apparent effect size of between 1 and 3
         points. At PbB levels above 1.2 µmol/litre (25 µg/dl), the
         relationship between PbB and IQ may differ. Estimates of effect
         size are group averages and only apply to the individual child in
         a probabilistic manner.

         Existing epidemiological studies do not provide definitive
         evidence of a threshold. Below the PbB range of 0.48-0.72 µmol/
         litre (10-15 µg/dl), the effects of confounding variables and
         limits in the precision in analytical and psychometric
         measurements increase the uncertainty attached to any estimate
         of effect. However, there is some evidence of an association
         below this range.

    b)   Animal studies provide support for a causal relationship between
         lead and nervous system effects, reporting deficits in cognitive
         functions at PbB levels as low as 0.53-0.72 µmol/litre
         (11-15 µg/dl) which can persist well beyond the termination of
         lead exposure.

    c)   Reduction in human peripheral nerve conduction velocity may occur
         with PbB levels as low as 1.44 µmol/litre (30 µg/dl). In
         addition, sensory motor function may be impaired with PbB levels
         as low as about 1.92 µmol/litre (40 µg/dl), and autonomic nervous
         system function (electrocardiographic R-R interval variability)
         may be affected at an average PbB level of approximately
         1.68 µmol/litre (35 µg/dl). The risk of lead nephropathy is
         increased in workers with PbB levels above 2.88 µmol/litre
         (60 µg/dl). However, recent studies using more sensitive
         indicators of renal function suggest renal effects at lower
         levels of lead exposure.

    d)   Lead exposure is associated with a small increase in blood
         pressure. The likely order of magnitude is that for any two-fold
         increase in PbB level (e.g., from 0.8 to 1.6 µmol/litre, i.e.
         16.6 to 33.3 µg/dl), there is a mean 1 mmHg increase in systolic
         blood pressure. The association with diastolic pressure is of a
         similar but smaller magnitude. However, there is doubt regarding
         whether these statistical associations are really due to an
         effect of lead exposure or are an artifact due to confounding
         factors.

    e)   Some but not all epidemiological studies show a dose dependent
         association of pre-term delivery and some indices of fetal growth
         and maturation at PbB levels of 0.72 µmol/litre (15 µg/dl) or
         more.

    f)   The evidence for carcinogenicity of lead and several inorganic
         lead compounds in humans is inadequate.

    g)   Effects of lead on a number of enzyme systems and biochemical
         parameters have been demonstrated. The PbB levels, above which
         effects are demonstrable with current techniques for the
         parameters that may have clinical significance, are all greater
         than 0.96 µmol/litre (20 µg/dl). Some effects on enzymes are
         demonstrable at lower PbB levels, but the clinical significance
         is uncertain.

    2.  IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES, AND ANALYTICAL METHODS

    2.1  Physical and chemical properties of lead and its compounds

         Lead (atomic number, 82; relative atomic mass, 207.19; specific
    gravity, 11.34) is a bluish or silvery grey soft metal. The melting
    point is 327.5°C and the boiling point at atmospheric pressure 1740°C.
    It has four naturally occurring isotopes (208, 206, 207, and 204 in
    order of abundance), but the isotopic ratios for various mineral
    sources may differ. This property has been exploited in
    non-radioactive-tracer environmental and metabolic studies. The
    physical and chemical properties of elemental lead and some lead
    compounds are summarized in Table 1.

         Although lead has four electrons in its valence shell, only two
    ionize readily. The usual oxidation state of lead in inorganic
    compounds is therefore +2 rather than +4. The inorganic salts of lead,
    such as lead sulfide and the oxides of lead, are generally poorly
    soluble in water. However, the nitrate, chlorate and, to a much lesser
    degree, the chloride are water soluble. Some of the salts formed with
    organic acids, e.g., lead oxalate, are also insoluble, but the acetate
    is relatively soluble, as shown in Table 1.

         Under appropriate conditions of synthesis, stable compounds are
    formed in which lead is directly bound to a carbon atom. Industrially
    synthesized lead-carbon compounds include tetraethyllead and
    tetramethyllead, which are of importance as fuel additives and, hence,
    are sources of environmental lead. 

    2.2  Analytical procedures

         In recent years substantial advances have been made in developing
    methods for the quantification of metals at low concentrations. In
    order to provide improved quality assurance of such measurements,
    various reference materials in different matrices have been produced
    (Muramatsu & Parr, 1985). To ensure adequate quality control, the
    analyst should choose a reference material that matches as closely as
    possible the experimental samples to be analysed. Choices are based
    upon matrix type and concentration of the element of interest. A
    summary of data on 60 biological and 40 environmental (non-biological)
    reference materials has been compiled by Muramatsu & Parr (1985).

         With the increased interest in measuring lead in the low µg/kg
    and µg/m3 range in both environmental and biological samples, there
    is need for particular attention to analytical sensitivity and
    reliability. As lower concentrations are measured, problems of
    laboratory contamination become more significant and quality control
    and quality assurance programmes are important. Because of these
    concerns, all analytical results for lead should report the laboratory


        Table 1.  Physical and chemical data on lead and selected lead compoundsa
                                                                                                                                              
    Name              Synonym         Relative atomic/   Melting point       Boiling point   Solubility in cold   Soluble in
                      and formula     molecular mass     (°C)                   (°C)         water (g/litre)
                                                                                                                                              

    Lead              Pb                  207.19         327.502                1740         insoluble            HNO3; hot concentrated
                                                                                                                  H2SO4; hot water;
                                                                                                                  glycerine; alcohol (slightly)

    Lead salts

      acetate         Pb(C2H3O2)2         325.28         280                    -            443

      carbonate       cerrusite PbCO3     267.20         315 (decomposes)                    0.0011               acid; alkali; decomposes in
                                                                                                                  hot water

      chlorate        Pb(ClO3)2           374.09         230 (decomposes)                    very soluble         alcohol

      chloride        cotunite PbCl2      278.10         501                    950          919                  NH4 salts; slightly in dilute
                                                                                                                  HCl and in NH3; hot water
                                                                                                                  (33.4 g/litre)

      nitrate         Pb(NO3)2            331.20         470 (decomposes)                    376.5                alcohol; alkali, NH3; hot
                                                                                                                  water (1270 g/litre)

      orthophosphate  Pb3(PO4)2           811.51         1014                                0.00014              alkali; HNO3

      oxalate         PbC2O4              295.21         300 (decomposes)                    0.0016               HNO3

      dioxide         plattnerite PbO2    239.19         290 (decomposes)                    insoluble            dilute HCl; acetic acid
                                                                                                                  (slightly)

      monoxide        litharge PbO        223.19         888                                 0.017                dilute HNO3; acetic acid
                                                                                                                                              

    Table 1 (cont'd)
                                                                                                                                              
    Name              Synonym         Relative atomic/   Melting point       Boiling point   Solubility in cold   Soluble in
                      and formula     molecular mass     (°C)                   (°C)         water (g/litre)
                                                                                                                                              

      sulfate         anglesite PbSO4     303.25         1170                                0.0425               NH4 salts; concentrated
                                                                                                                  H2SO4 (slightly)

      sulfide         galena PbS          239.25         1114                                0.00086              acid
                                                                                                                                              

    a  Data from Weast (1985)
    

    performance for reference standards and for parallel blank
    measurements of sample contamination for the entire analytical
    process. Without these, the validity of the data should be questioned.

    2.2.1  Sampling procedures

         Particular attention should be paid to the cleanliness of
    equipment and glassware and the purity of the chemicals to prevent
    secondary contamination by lead.

         For the collection of samples, standard trace element methods are
    generally required (Behne, 1980) with adequate quality control
    procedures (Friberg, 1988; Jorhem & Slorach, 1988, Vahter & Friberg,
    1988). Quality control samples for blood, faeces, air filters and dust
    have been described (Lind et al., 1988).

    2.2.1.1  Sampling of environmental media

         In air sampling, both high-volume samplers and low-volume
    techniques have been used. It should be noted that the collection
    characteristics of high-volume samplers are strongly affected by
    particle size and the orientation of the sampler. For particles larger
    than 5 µm in diameter the high-volume sampler system is unlikely to
    collect representative samples (US EPA, 1986a). As in all sampling for
    suspended particulate matter, the accuracy of volume meters should be
    checked periodically. The size of the pores in filters for collecting
    lead-containing particles should be small, possibly less than 0.2 µm
    for glass-fibre filters (Lee & Goranson, 1972).

         Depending on the purpose of sampling, care should be taken to
    select the appropriate site for sampling devices and to achieve the
    best possible sampling conditions by:

    *    estimating the amount of particulate required for analysis before
         deciding on the sample volume and the sampling procedure;

    *    placing the sampling devices in the appropriate position (e.g.,
         in the breathing zone, level with inlet tubes of house
         ventilators, at window level in the case of a traffic-laden town
         street, at a reasonable distance from the highway in uninhabited
         zones, etc);

    *    taking the samples at appropriate rates and volumes (e.g., daily
         breathing volumes, daily ventilating capacities of installations)
         and for a sufficient time to make possible the estimation of the
         average concentration (e.g., during a work-shift, or a 24-h or
         longer period for general population exposure);

    *    taking into account the use of areas under study (cattle grazing,
         recreational zones, children's playgrounds, etc).

         In addition, whenever possible a procedure should be used that
    makes it possible to evaluate particle-size distribution and the
    physicochemical properties of the lead compounds involved, including
    the shape of the particles and the state of their aggregation.

         Lead may be found in water bound to particulate matter as soluble
    complexes or soluble compounds. Techniques for sampling water must
    take this into account. It is necessary to sample water without
    fractionation (filtration) when total lead levels are required.
    Because of the potential for metals from low ionic strength waters to
    be adsorbed onto the surfaces of some containers, samples should be
    acidified (US EPA, 1986a). Selection, cleaning, and conditioning of
    storage and sample containers deserve special attention (Moody, 1982).

         The preparation of soil and dust samples for lead analyses
    usually involves drying (at 100°C), homogenization by grinding, and
    sieving (Thornton & Webb 1975; Bolter et al., 1975). Brown & Black
    (1983) have discussed the issues related to quality assurance and
    quality control in the collection and analysis of soil samples. Most
    reports of lead in soil provide the total elemental abundance either
    by acid extraction or X-ray fluorescence. However, the leachable or
    bioavailable fraction is of special interest.

         For the study of the dietary intake of lead from food, two
    general methods have been utilized. The advantages and disadvantages
    of the "duplicate portions" technique and the equivalent composite
    technique ("market basket") have been reviewed by Pekkarinen (1970).
    Although the duplicate portions (duplicate diets) technique can define
    variability in consumption, it is expensive, and the sampling and
    analytical procedures involved are complicated and limit the number of
    individuals included in any study. With the equivalent composite
    technique, the economy and ease of collection must be considered in
    the light of the variability of results obtained due to uncertainties
    in knowledge of actual preparation techniques, including possible lead
    levels in water used for processing in individual homes.

         The quantity of lead likely to be leached from ceramic surfaces
    by different foods and beverages may be assessed using dilute acetic
    acid solutions (1 to 4%) at temperatures in the range 20 to 100°C for
    times ranging from 30 min to more than 24 h (Laurs, 1976; Merwin,
    1976).

         Colorimetric methods are suitable for screening inorganic
    materials such as pottery or paint for lead. Positive reactions
    require confirmation by established quantitative methods. Spot tests
    using dithizone, rhodizonate and iodide (Feigl et al., 1972) are
    available.

    2.2.1.2  Sampling of biological materials

         The main problem in the sampling of body fluids and tissues for
    lead analysis is potential secondary contamination with lead. The low
    general population blood lead (PbB) levels in many regions of the
    world are complicating screening efforts, requiring levels of
    analytical precision and sensitivity that can be achieved only through
    intensive QA/QC programmes. Issues related to such sampling have been
    examined in detail by US EPA (1986a).

         Special precautions are needed to ensure that all venous
    blood-collecting and blood-storage materials are as free from lead as
    possible (IPCS, 1977). All glass equipment involved in blood
    collection and storage should be made of lead-free silicate glass,
    rinsed first in mineral acid, then with copious amounts of
    glass-distilled or deionized water. Polypropylene syringes have been
    recommended (NAS-NRC, l972). Needles should be of stainless steel with
    polypropylene hubs. Blood is often drawn directly from the needle into
    vacuum tubes. It is wise to confirm periodically the absence of
    significant amounts of lead in the anticoagulant used in the blood
    container as well as monitoring the contamination level (blank) for
    the entire analytical process.

         New analytical techniques make it possible to determine lead
    concentrations in microlitre quantities of blood. The trend towards
    the procurement of micro-samples of blood by skin prick increases the
    chance of secondary contamination of the blood. Systematic
    investigation on the significance of this problem has been reported
    (Mitchell et al., 1974; Mahaffey et al., 1979; DeSilva & Donnan,
    1980). Mitchell et al. (l974) describe a procedure whereby sample
    contamination can be reduced by spraying collodion over the cleansed
    skin before lancing. The correlation between the concentration of lead
    in micro-samples and in macro-samples obtained by venepuncture was
    fairly good (r=0.92) over a wide range of PbB concentrations
    (0.48-4.41 µmol/litre or 10-92 µg/dl whole blood). Mahaffey et al.,
    (1979b) found that capillary blood levels in a comparison test were
    systematically higher than corresponding venous blood levels; similar
    elevations have been reported by DeSilva & Donnan (1980). Since about
    1980 the requirement for reliable and accurate micro procedures has
    resulted in the development of good protocols. Sinclair & Dohnt (1984)
    described a procedure which resulted in the ability to collect
    capillary samples with PbB levels only 3.3% higher than the presumably
    correct venous value. This procedure has been used in the Port Pirie
    Cohort Study (Baghurst et al., 1985, 1992) and for routine
    surveillance in the Port Pirie Lead Decontamination Program (Calder et
    al., 1990). Also, Lyngbye et al. (1990b) have shown that capillary
    sampling without lead contamination is possible. Routine validation by
    cross-comparison with venous blood samples should be undertaken on a
    regular basis.

         The same general precautions to avoid contamination must be taken
    in the collection of urine samples as in the collection of blood
    samples. Additionally, special care must be taken to prevent
    precipitation during storage.

    2.2.2  Analytical methods for lead

         A number of analytical methods exist for determination of lead in
    environmental and biological samples. These methods differ enormously
    in their costs (e.g., sophisticated equipment, an adequate
    infrastructure to maintain laboratory conditions and chemical
    supplies) and personnel requirements (e.g., availability of skilled
    personnel in adequate numbers for the work to be undertaken). Both
    accuracy and precision of any of the methods can be affected greatly
    by contamination of samples within the laboratory. It is important to
    utilize the principles of a "clean" laboratory described by Patterson
    & Settle (1976) and Everson & Patterson (1980).

         It is not the purpose of this section to provide an exhaustive
    description of the analytical methods that could be available to
    detect and quantify lead levels in environmental and biological
    samples. However, an attempt will be made to identify well-established
    methods in current use and to provide information on their application
    to assist in the interpretation of experimental and epidemiological
    studies.

    2.2.2.1  Analysis of lead in environmental samples

         The most common methods used for the analysis of lead in samples
    from air, water, dust, sediment, soil and foodstuffs are flame atomic
    absorption spectrometry (AAS), graphite furnace atomic absorption
    spectrometry (GFAAS), anodic stripping voltametry (ASV), inductively
    coupled plasma-atomic emission spectroscopy (ICP-AES), and X-ray
    fluorescence spectroscopy (XRFS). The reference method for the
    determination of the absolute amounts of lead is by isotope dilution
    mass spectrometry (IDMS) (Settle & Patterson, 1980; Grandjean & Olsen,
    1984; US EPA, 1986a), but due to equipment costs and required
    expertise, it is not widely used. Spectrophotometric methods, using
    diphenylthiocarbazone as the colorimetric reagent, were widely used in
    the past; they are less sensitive and are labour-intensive but are
    still appropriate. The advantages and disadvantages were described by
    Skogerboe et al. (1977).

         Gould et al. (1988) utilized a citric acid solution on filter
    paper to leach lead from glazed ceramic and/or enamelled metal-ware.
    When treated with a lead-sensitive chromogen, there is a reaction
    indicating the presence of lead on the paper. The minimal amount of
    lead required to produce an observable reaction was 0.25 µg/cm2; the
    maximum amount tested was 5 µg/cm2. A colorimetric test based on the
    use of sodium sulfide in solution is used to estimate lead in paint

    films. It is possible to determine lead concentrations greater than
    1 mg/cm2 of dried paint 90% of the time when the method is used by a
    trained chemical laboratory technician.

         Table 2 summarizes the utility of several representative methods
    for specific environmental media.

    2.2.2.2  Analysis of lead in biological materials

         Biological samples present special problems for the analyst
    because of the low lead concentrations and matrix effects. Most
    analytical techniques developed to detect and quantify lead can be
    adapted to the analysis of such biological materials as blood, urine,
    serum, cerebrospinal fluid, solid tissues, hair, teeth and bone.
    However, certain techniques are more often used for specific matrices.

         Currently, the most commonly used methods are AAS, GFAAS, ASV,
    and ICP-AES. Spectrophotometric methods were commonly used in the past
    and can be useful. Other specialized methods for lead analysis are
    XRFS, neutron activation analysis (NAA), inductively coupled
    plasma-mass spectrometry (ICP-MS), and IDMS. Table 3 summarizes the
    utility of several analytical procedures applied to various biological
    matrices. Included in this table are examples of the application of
    XRFS (Christoffersson et al., 1986; Wielopolski et al., 1986; Nilsson
    et al., 1991) for the determination  in situ of the body burden of
    lead.

    2.2.2.3  Analytical procedures for biomarkers of lead exposure and
             effect

         Using standard clinical laboratory techniques, analytical
    procedures have been developed: delta-aminolaevulinic acid (ALA);
    delta-aminolaevulinic acid dehydratase (ALAD); urinary coproporphyrin
    (CPU) and erythrocyte protoporphyrin (EP). All of these assays are
    well established and reliable (Grandjean & Olsen 1984; US EPA, 1986a).
    These biochemical parameters are influenced by physiological factors
    other than lead. They lack the specificity and sensitivity of PbB
    measurements as an index of either current lead exposures or body
    stores of lead.

    2.3  Conversion factors

         1 µg/dl = 0.048 µmol/litre
         1 µmol/litre = 20.7 µg/dl

         Using the above conversion factor, blood lead concentrations are
    given as µmol/litre with the equivalent µg/dl in brackets. Calculated
    figures have not been rounded and added precision is not to be
    inferred from the number of significant figures.


        Table 2.  Analytical methods for determining lead in environmental samplesa
                                                                                                                                               
    Sample type       Preparation method                             Analytical method          Sample detection   Percentage   Reference
                                                                                                limit              recovery
                                                                                                                                               

    Air               collect particulate matter on membrane         ASV with mercury-graphite  0.16 µg/m3         90-110       NIOSH (1977b)
     (particulate     filter; wet ash with HNO3/HClO4/H2SO4;         electrode (NIOSH method
     lead)            dissolve in acetate buffer                     P&CAM 191)

    Air               collect particulate matter on cellulose        ICP-AES (NIOSH method      0.34 µg/m3         95-105       NIOSH (1981)
     (particulate     acetate filter; wet ash with HNO3/HClO4        P&CAM 351)
     lead)

    Air               collect particulate matter on filter;          AAS                        0.1 µg/m3          93           Scott et al.
     (particulate     dry ash; extract with HNO3/HCl; dilute         AES                        0.15 µg/m3         102          (1976)
     lead)            with HNO3

    Air               sample on cellulose acetate filter;            AAS                        8 ng/litre         100-101      Nerin et al.
     (particulate     dissolve in HNO3 with heat; add HCl/H2O2                                                                  (1989)
     lead)            and react in hydride generator with sodium
                      borohydride to generate lead hydride

    Air               collect sample on filter; spike filter with    IDMS                       0.1 ng/m3          NR           Volkening et
     (particulate     206Pb; dissolve filter in NaOH; acidify;                                                                  al. (1988)
     lead)            separate lead by electrodeposition; dissolve
                      in acid

    Water             digest sample with acid; heat; dilute with     AAS                        1.0 ng/g           NR           Chau et al.
     (total lead)     water                                                                                                     (1979)

    Soil              dry sample and sieve for XRF; digest sieved    XRF                        NR                 65-98        Krueger &
                      sample with HNO3 and heat for AAS              AAS                        NR                 63-68        Duguay (1989)

    Soil              dry sample, dry ash; digest with acid          AAS                        2 µg/g             79-103       Beyer &
                      and dilute with water                                                                                     Cromartie (1987)
                                                                                                                                               

    Table 2 (cont'd)
                                                                                                                                               
    Sample type       Preparation method                             Analytical method          Sample detection   Percentage   Reference
                                                                                                limit              recovery
                                                                                                                                               

    Soil, waste,      digest sample with acid; dilute with water     AAS (EPA method 7420)      0.1 mg/litre       NR           US EPA (1986b)
     and ground       and filter                                     GFAAS (EPA method 7421)    1 µg/litre         NR
     water

    Soil, dust        digest sample with hot acid; dry; redissolve   AAS                        12 ng/g            > 80         Que Hee et al.
     and paint        in HNO3                                                                                                   (1985b)

    Sediment, fish,   digest sample with acid; heat; dilute with     AAS                        50 ng/g            NR           Chau et al.
     vegetation       water                                                                     (sediment)                      (1980)
     (total lead)                                                                               10 ng/g (fish      NR
                                                                                                and vegetation)

    Milk              add 50 µl (C2H5)4NOH in ethanol to 25 µl       GFAAS                      NR                 NR           Michaelson &
                      milk; heat and dilute with water to 125 µl                                                                Sauerhoff (1974)

    Evaporated        dry ash sample; dissolve in HNO3               ASV                        0.005 µg/g         99           Capar & Rigsby
     milk                                                                                                                       (1989)

    Agricultural      dry ash sample with H2SO4 and HNO3;            DPASV                      0.4 ng/g           85-106       Satzger et al.
     crops            dilute with water                                                                                         (1982)

    Grains, milk,     bomb digest sample with acid; heat or          GFAAS                      20 µg/g (bomb)     85-107       Ellen & Van
     mussels, fish    digest with acid and dry ash; dissolve                                    5 µg/g (dry ash)   75-107       Loon (1990)
                      in acid; dilute with water                     DPASV                      NR                 82-120
                                                                                                                                               

    Table 2 (cont'd)
                                                                                                                                               
    Sample type       Preparation method                             Analytical method          Sample detection   Percentage   Reference
                                                                                                limit              recovery
                                                                                                                                               

    Citrus leaves     chop or pulverize sample; digest with hot      ICP-AES                    10-50 µg/litre     75-82        Que Hee &
     and paint        acid; dry; redissolve in acid                                                                (citrus      Boyle (1988)
                                                                                                                   leaves)
                                                                                                                   89-96
                                                                                                                   (paint)

                                                                                                                                               

    a  AAS = atomic absorption; AES = atomic emissions spectroscopy; ASV = anode stripping voltametry; (C2H5)4NOH = tetraethylammonium
       hydroxide; DPASV = differential pulse anodic stripping voltametry; EPA = US Environmental Protection Agency; GFAAS = graphite furnace
       atomic absorption spectrometry; HCl = hydrochloric acid; HClO4 = perchloric acid; HNO3 = nitric acid; H2O2 = hydrogen peroxide;
       H2SO4 = sulfuric acid; ICP-AES = inductively coupled plasma/atomic emission spectroscopy; IDMS = isotope dilution mass spectrometry;
       NaOH = sodium hydroxide; NIOSH = National Institute for Occupational Safety and Health; NR = not reported; XRF = X-ray fluorescence

    Table 3.  Analytical methods for determining lead in biological materialsa
                                                                                                                                               
    Sample      Preparation method                           Analytical method            Sample detection         Percentage    Reference
    type                                                                                  limit                    recovery
                                                                                                                                               

    Blood       wet ash sample with acid mixtures;           ASV with mercury-graphite    0.192 µmol/litre         95-105        NIOSH (1977c)
                dissolve residue in dilute HClO4             electrode (NIOSH method      (4 µg/dl)
                                                             P&CAM 195)

    Blood       wet ash sample with HNO3; dissolve           GFAAS (NIOSH method          0.48 µmol/litre          NR            NIOSH (1977e)
                residue in dilute HNO3                       P&CAM 214)                   (10 µg/dl)

    Blood       dilute sample with Triton X-100(R); add      GFAAS                        0.011 µmol/litre         93-105        Aguilera de
                nitric acid and diammonium phosphate                                      (0.24 µg/dl)                           Benzo et al.
                                                                                                                                 (1989)

    Blood       dilute sample with ammonia solution          ICP-MS                       0.072 µmol/litre         96-111        Delves &
                containing Triton X-100(R); analyse                                       (1.5 µg/dl)                            Campbell
                                                                                                                                 (1988)

    Blood       dilute sample in 0.2% Triton X-100(R)        GFAAS                        approx. 0.072            97-150        Que Hee et
                and water; analyse                                                        µmol/litre (approx.                    al. (1985a)
                                                                                          1.5 µg/dl)

    Blood and   wet ash sample with HNO3, complex            Spectrophotometry            0.144 µmol/litre         97            NIOSH (1977a)
     urine      with dephenylthiocarbazone and               (NIOSH method                (3.0 µg/dl) (blood);
                extract with chloroform                      P&CAM 102)                   0.0576 µmol/litre        97
                                                                                          (12 µg/litre) (urine)

    Serum,      filter sample if needed; dilute with         ICP-AES                      0.048-0.240 µmol/litre   85 (serum)    Que Hee &
     blood and  acid or water                                                             (1.0-5.0 µg/dl)                        Boyle (1988)
     urine

    Urine       wet ash sample with acid mixture and         ASV with mercury-graphite    0.0192 µmol/litre        90-110        NIOSH (1977d)
                dissolve in dilute HClO4                     electrode (NIOSH method      (4 µg/litre)
                                                             P&CAM 200)
                                                                                                                                               

    Table 3 (cont'd)
                                                                                                                                               

    Sample      Preparation method                           Analytical method            Sample detection         Percentage    Reference
    type                                                                                  limit                    recovery
                                                                                                                                               

    Liver,      bomb digest sample with acid and heat,       GFAAS                        20 µg/g (bomb);          85-107        Ellen & Van
     kidney,    or digest with acid and dry ash; dissolve                                                          (bomb);       Loon (1990)
     muscle     in acid; dilute with water                                                5 µg/g (dry ashing)      75-107 (dry
                                                                                                                   ashing)

    Bone        direct partially polarized photons at        XRF                          20 µg/g                  NR            Christoffersson
                second phalanx of left forefinger                                                                                et al. (1986)
                (non-invasive technique)

    Bone        direct partially polarized photons at        XRF                          20 µg/g                  NR            Wielopolski
                anteromedial skin surface of mid-tibia                                                                           et al. (1986)
                (non-invasive technique)

    Teeth       clean and section tooth; digest with         ASV                          NR                       83-114        Rabinowitz
                HNO3; evaporate; redissolve in buffer                                                                            et al. (1989)
                solution

    Teeth       dry ash sample; crush; dry ash again;        AAS                          NR                       90-110        Steenhout &
                dissolve in HNO3                                                                                                 Pourtois
                                                                                                                                 (1981)
                                                                                                                                               

    a  AAS = atomic absorption spectrometry; ASV = anode stripping voltametry; GFAAS = graphite furnace atomic absorption spectrometry;
       HClO4 = perchloric acid; HNO3 = nitric acid; ICP-AES = inductively coupled plasma-atomic emission spectroscopy;
       ICP-MS = inductively coupled plasma-mass spectrometry; NIOSH = National Institute for Occupational Safety and Health;
       NR = not reported; XRF = X-ray fluorescence
    

    3.  SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE

    3.1  Natural occurrence

         Because lead is relatively abundant in the earth's crust it is
    found naturally throughout the world. The major natural sources of
    lead are volcanic emissions, geochemical weathering, and emissions
    from sea spray. A small amount of radioisotopic lead (207Pb) is
    derived from the decay of radon gas released from geological sources.
    It has been estimated that the worldwide natural emission rates of
    lead are of the order of 19 000 tonnes/year (Nriagu & Pacyna, 1988),
    with volcanic sources accounting for 6400 tonnes/year (Nriagu, 1979).

         Owing to centuries of human exploitation of lead resources, it is
    difficult to determine the natural content of lead in most ecosystems.
    Data on environmental levels, uses, and sources of lead have been
    summarized in a recent review (OECD, 1993).

    3.1.1  Rocks and soils

         The average concentration of lead in the earth's crust is between
    10 and 20 mg/kg (IPCS, 1989). The major geological sources of lead are
    in igneous and metamorphic rocks.

         The soil is the most important repository in terrestrial
    ecosystems for contaminants of anthropogenic origin (Nriagu & Pacyna,
    1988; Nriagu, 1989). The lead content of soils (which are for
    discussion purposes distinguished here from surface dusts) is greatly
    influenced by anthropogenic activities and by long- and short-range
    airborne transport of lead from various sources. Both dry and wet
    deposition are important routes of input.

         Lead in soil may be relatively insoluble (as a sulfate, carbonate
    or oxide), soluble, adsorbed onto clays, adsorbed and coprecipitated
    with sesquioxides, adsorbed onto colloidal organic matter, or
    complexed with organic moieties in soil (US EPA, 1986a; IPCS, 1989).
    Soil pH, content of humic and fulvic acids, and amount of organic
    matter influence the content and mobility of lead in soils. Since
    acidic conditions favour the solubilization and leaching of lead from
    the solid phase, acidic soils tend to have lower lead concentrations
    when analysed as dry soil. Humic and fulvic acids can also mobilize
    lead, and certain complex organic molecules can act as chelators of
    lead (IPCS, 1989).

         Background levels of lead in soil are in the range of 10-70 mg/kg
    (GEMS, 1985). Similar results have been found in studies of mobile
    source contamination near highways; soil taken at distances of
    50-100 m from highways (outside the range of immediate impact from
    traffic emissions) usually shows levels of lead below 40 mg/kg. In the
    1985, GEMS survey of selected countries, lead concentrations in
    topsoil from Malta were found to have a mean of 54 mg/kg in areas at

    least 5 m from roadways; less than one metre from roadways the mean
    concentration was 138 mg/kg. A 1977 report from Sweden found a mean of
    16 mg/kg in non-contaminated areas (GEMS, 1985).

    3.1.2  Sediments

         Sediments from freshwater and marine environments have been
    studied for lead content. This compartment provides a unique record of
    the history of changes in global lead fluxes (Patterson, 1983). Levels
    of lead in sediments dated before the onset of the industrial
    revolution in Western Europe show very low levels, less than 10% of
    current levels (Flegal et al., 1987). The average background level of
    lead in marine sediments off southern California was reported by
    Flegal et al. (1987) to be 1.3 mg/kg.

    3.1.3  Water

         Flegal et al. (1987) estimate that the natural concentration of
    lead in surface water is about 0.02 µg/litre. In general, lead is not
    found in ground or surface waters at concentrations above 10 µg/litre
    (IPCS, 1989).

         Data from oceans indicate very low levels of lead in sea-water
    samples not affected directly by significant sources of lead. Water
    samples taken from an area of the Pacific, where annual windborne-
    input fluxes of lead are estimated to be 3 mg/cm2, have lead
    concentrations of 3.5 ng/litre (0-100 m depth) and 0.9 ng/litre at
    depths greater than 2500 m. In contrast, water samples taken from the
    north Atlantic, where annual windborne-input fluxes of lead are
    170 mg/cm2, contain 34 ng lead/litre at the surface and 5 ng/litre
    in depths below 2500 m (Patterson, 1983). Settle & Patterson (1980)
    have estimated that prehistoric oceans contained 0.5 ng/litre lead.
    Flegal et al. (1987) have estimated that over 95% of the lead in
    off-shore surface waters is the result of windborne inputs. However,
    in coastal waters near Monterey (California, USA), higher
    concentrations of lead were found in sea water, sediments and
    organisms; these elevations were related to specific sources by
    systematic isotope analyses (Flegal et al., 1987).

    3.1.4  Air

         Anthropogenic inputs of lead from a range of sources have
    resulted in global dispersion of both inorganic and organic species of
    lead into the air, of which 80-90% is derived from alkyllead fuel
    additives (WHO, 1987). Nriagu & Pacyna (1988) estimated that a total
    of 330 000 tonnes of lead is discharged directly into the atmosphere
    each year. Estimations of pre-industrial levels of lead in air from
    natural origins (volcanic emissions, crustal weathering, radon decay
    and sea-spray releases) are in the range of 0.01-0.1 µg/m3 (US NRC,
    1980). The lowest level reported since 1975 is 0.076 ng/m3 measured
    at the South Pole (US EPA, 1986a).

    3.1.5  Plants

         Lead occurs naturally in plants and results from both deposition
    and uptake. There is a positive linear relationship between lead
    concentrations in plants and soil (Davies & Thornton, 1989). As with
    other environmental compartments, measurement of "background" levels
    of lead in plants is complicated by the general contamination of the
    globe from centuries of lead use, which has included direct
    application of lead-containing chemicals in agriculture (see below)
    and contamination of fertilizers with lead. Lead has been measured in
    superphosphate fertilizer at concentrations as high as 92 mg/kg (Lisk,
    1972). Sewage sludge, used as a source of nutrients in agriculture,
    may contain even higher levels of lead. The concentration of lead in
    sewage sludge is typically < 1000 mg/kg. Levels as high as 26 g/kg
    have been measured in the USA (Chaney et al., 1984). Soil receiving
    heavy sludge applications over long periods of time (years) contained
    425 mg lead/kg; the concentration in untreated soil was 47 mg/kg
    (Beckett, 1979).

    3.1.6  Environmental contamination from natural sources

         The contribution of natural sources of lead to human exposure is
    small. As a result of various breakdown processes, rocks yield lead
    which is transferred to the biosphere and the atmosphere, and,
    ultimately, back to the earth's crust in the form of sedimentary
    rocks. Soluble lead has for thousands of years entered the oceans with
    river discharges, and the rate has been estimated by Patterson (1965)
    to be around 17 000 tonnes/year. Sources contributing to airborne lead
    are silicate dusts, volcanic halogen aerosols, forest fires, sea salts
    aerosols, meteoric and meteorite residues, and lead derived from the
    decay of radon. While the lead content of most coals is relatively
    low, coal fly ash is enriched in lead (Hutton et al., 1988) and is a
    source of environmental contamination.

    3.2  Anthropogenic sources

         World lead consumption has steadily increased over the period
    1965-1990 and was about 5.6 × 106 tonnes in 1990 (OECD, 1993).

         Further review of the data summarized by OECD (1993) indicates a
    change in consumption patterns worldwide. Although the consumption of
    lead within the 24 countries of the OECD increased only slightly over
    the decade from 1980 to 1990, consumption within less developed
    economies (Africa and Asia) increased from 315 000 tonnes in 1970 to
    844 000 tonnes in 1990.

    3.2.1  Lead mining

         Lead occurs in a variety of minerals, the most important of which
    are galena (PbS), cerrusite (PbCO3) and anglesite (PbSO4). Galena
    is by far the most important source of primary lead. It occurs mostly

    in deposits associated with other minerals, particularly those
    containing zinc. Mixed lead and zinc ores account for about 70% of
    total primary lead supplies. Ores containing mainly lead account for
    about 20% and the remaining 10% is obtained as a by-product from other
    deposits, such as mixed copper-zinc deposits. The proportions of
    various metals may differ in the ores of different countries. Silver
    is the most important of the other metals frequently present in lead
    deposits, but copper may also be present in concentrations high enough
    to be commercially important. Other minor constituents of lead ores
    are gold, bismuth, antimony, arsenic, cadmium, tin, gallium, thallium,
    indium, germanium and tellurium.

         The major countries producing lead from mining activity during
    1987-1991 were the USA, Canada, Australia, Peru, the former USSR and
    Mexico, as shown in Table 4. Other countries producing lead from lead
    ores include China, the former Yugoslavia, Morocco, Spain, Sweden and
    Tunisia. In general, the level of world production of lead from mining
    activities has remained relatively constant at about 3.3 × 106
    tonnes between 1988 and 1991 (ILZSG, 1992); this represents roughly
    60% of the world demand for lead.


    Table 4.  Major countries producing lead from ore and ore
              concentratesa
                                                                         

    Country      1987        1988        1989        1990        1991
                                                                         

    Canada       423 200     366 600     276 100     241 300     278 100

    USA          318 300     395 700     419 300     495 200     483 300

    Ex-USSR      510 000     520 000     500 000     490 000       --

    Australia    489 200     462 000     495 000     570 000     579 000

    Mexico       177 200     178 100     163 000     174 100     158 800

    Peru         204 000     149 000     192 200     187 800     199 100
                                                                         

    a  From: World Bureau of Metal Statistics (1992)


    3.2.2  Smelting and refining

         Smelting and refining are classified as either primary or
    secondary, the former producing refined lead products from ores or
    concentrates (primary lead) and the latter producing lead by
    recovering it from lead-bearing scrap and waste materials (secondary
    lead). Secondary lead is derived from processing what is termed new

    scrap arising during manufacturing processes and recycled old scrap
    arising from waste materials containing lead. Most scrap is from old
    sources, of which the most important are lead plates from batteries,
    solder, common babbitt, soft lead, lead solders, cable coverings, type
    metals, dross and other lead-containing products. There has been an
    increasing contribution of secondary lead sources to the total
    worldwide production of lead, as shown in Table 5 (World Bureau of
    Metal Statistics, 1992). Secondary sources of lead supplied between 35
    and 40% of world production during the period from 1970 to 1990.

    3.2.3  Environmental pollution from production of lead

         Mining operations and the smelting and refining of both primary
    and secondary lead are known to cause contamination of the nearby
    environment. The nature and extent of contamination depends on many
    factors, including the level of production, the effectiveness of
    emission controls, climate, topography and other local factors.
    Concentrations are usually highest within 3 km of the point source (US
    EPA, 1989). A report from China found that lead levels in ambient air,
    plants and soil increased proportionally with proximity to a large
    primary smelter; at 50 m from the source, the air lead level was
    60 µg/m3, the lead level in plants was 29.1 mg/kg, and soil lead
    level was 170 mg/kg (Wang, 1984). However, some earlier studies have
    shown air pollution and soil contamination as far as 10 km from
    smelters (Djuric et al., 1971; Kerin, 1973; Landrigan et al., 1975).


    Table 5.  Relative contribution of primary and secondary sources
              relative to world lead production (1987-1990)a
                                                                         

                        1987         1988          1989          1990
                                                                         

    Primary           422 100     3 414 200      3 286 500     3 324 500
    Secondary       2 045 600     2 103 900      2 272 900     2 254 800
                                                                         

    a  From: World Bureau of Metal Statistics (1992)


         The impacts of lead mining and smelting can persist for long
    periods of time. A study conducted in Wales, United Kingdom, in an
    area where lead mining began 2000 years ago and ended in the middle of
    the 20th century, found high concentrations of lead in soils (Davies
    et al., 1985). In Port Pirie, Australia, a community with one of the
    world's largest and oldest primary lead smelters, lead levels in soils
    were found to be grossly elevated, and the incidence of elevated blood
    lead levels in pregnant women and young children was also increased
    above that found in other communities in Australia (Wilson et al.,
    1986).

    3.3  Consumption and uses of lead and its compounds

         Lead has a combination of physical and chemical properties that
    have made it extremely useful industrially, i.e. high density, high
    opacity to gamma and X-ray energies, low sound conductance, a low
    melting point, exceptional malleability, high corrosion resistance,
    and stability. In 1990, 5.627 × 106 tonnes of lead were consumed
    worldwide (ILZSG, 1992). The twenty-four industrialized countries of
    the OECD consumed approximately 65% of this amount, with eastern
    Europe and the former USSR using 21%. Asia now utilizes about 9% of
    the world's lead production.

         The use patterns of refined lead vary from country to country.
    The situation in 1990 in Mexico is shown in Table 6, while end-use
    categories within OECD countries are summarized in Fig. 1, which
    indicates the changes between 1970 and 1990 (OECD, 1993).


    Table 6.  Principal uses of refined lead in Mexicoa
                                             

    Type of product      1988          1990
                         (%)           (%)
                                             

    Oxides               69.7          56.7
    Batteries             9.2          17.9
    Tetraethyllead        7.9          11.9b
    Cables                4.0           1.5
    Others                9.2          11.9
                                             

    a  ILZSG (1992)
    b  This does not reflect the introduction of lead-free petrol
       in 1990.

         From Fig. 1, it is evident that the largest use of lead within
    OECD countries is for battery production, whereas there has been a
    large drop in the demand for lead-containing gasoline additives.
    However, this pattern is not valid worldwide, e.g., concentrations in
    petrol range from zero in such countries as Japan and Thailand to
    1.12 g/litre in the Virgin Islands (Octel, 1991).

         In the past the use of lead in the chemical industry for the
    preparation of paints, pigments and coloured inks was widespread. Many
    countries have now restricted this use, and concentrations of lead
    greater than 0.06% (USA) and 0.5% (New Zealand) are not permitted in
    indoor paints (Albert & Badillo, 1991; OECD, 1993). In 1982, data from
    the United Kingdom (UK, 1982) indicated levels of lead between 2500
    and 3000 mg/kg in decorative glass paints and up to 448 g/kg in
    white-lead primer. Red-leadcontaining paints, still used widely to
    paint structural steel works, can contain up to 661 g lead/kg.

    FIGURE 1

         Other disperse uses of lead include lead solders (now banned in
    USA for use in drinking-water systems), ammunition (Novotny et al.,
    1987), foil on wine bottles (Wai et al., 1979) and cosmetics and
    folk-medicines (surma in Asia, Kohl in India, and Al Kohl in Saudi
    Arabia and Kuwait) (Fernando et al., 1981).

    3.4  Sources of environmental exposure

         As noted above, lead is a ubiquitous pollutant in the global
    ecosystem, as well as occurring naturally. Its uses have resulted in
    increases in soil, water and air lead levels to one to two orders of
    magnitude above those estimated to have prevailed prior to rapid
    industrialization in the 18th and 19th centuries (Patterson, 1983).
    Whereas in specific areas point sources may contribute significant
    amounts of lead to the environment, on a global scale, the combustion
    of alkyllead in petrol is the predominant source of increased lead in
    all compartments of the environment. This has been hypothesized based
    upon mass balance studies (Nriagu, 1979) and confirmed by the changes
    in environmental lead levels which have followed the significant
    reductions in worldwide use of alkyllead as a gasoline additive since
    the mid-1980s. For example, lead concentrations in Greenland snow
    decreased by a factor of 7.5 over a 20-year period from the late 1960s
    (Boutron et al., 1991).

         Nriagu & Pacyna (1988) have estimated the global emissions of
    lead to the atmosphere resulting from anthropogenic uses (Table 7).
    Current estimates (OECD, 1993) of emissions from mobile sources would
    be about 30% of the 1983 amounts. Similarly estimates of emissions of
    lead to soil in 1983 were made by Nriagu & Pacyna (1988) (Table 8).
    Since lead is never degraded, all lead which is shifted from
    geological sources by human technology eventually enters the
    environment through disposal, although this can be slowed by recycling
    and recovery.

         Municipal solid waste (MSW), solid waste, hazardous waste, sewage
    sludge, and industrial waste-water discharges all may contain lead at
    concentrations as high as 50 g/kg. Although few measurements of
    environmental lead concentrations in the vicinity of disposal sites
    have been conducted, analyses of fly and bottom ash from municipal
    incinerators show high concentrations (up to 50 g/kg) of lead (Wadge &
    Hutton, 1987), and land disposal sites which have received incinerator
    ash for a number of years show high levels of lead in soil (Hutton et
    al., 1988).

         Dusting and flaking of lead paint from surfaces can be a source
    of lead contamination in surface dust and soil near houses or
    buildings as well as contributing to the concentrations of lead in
    household dust. This process is a function of the type of paint and
    the age and state of repair of the structure. When lead paint is
    present on structures, both interior and exterior dusts have higher

    concentrations than otherwise would be expected (Thornton et al.,
    1985). Abatement of lead paint may be a major local source of
    environmental contamination, as shown by studies near school buildings
    in London (Rundle & Duggan, 1986). Removal of lead-based paints from
    bridges and water towers using improper techniques can also result in
    significant environmental contamination. Direct application of
    lead-contaminated sludge as fertilizers, and residues of lead arsenate
    from use in agriculture can lead to the contamination of soil, surface
    water and ground water. In local aquatic environments, pollution can
    result from leaching of lead from lead shot, shotgun cartridges and
    fishing weights (IPCS, 1989). Coal contains small amounts of lead,
    which can be concentrated in fly ash from coal combustion (Wadge &
    Hutton, 1987) or in stack emissions (Table 8).

        Table 7.  Estimated worldwide anthropogenic emissions of lead to
              the atmosphere (1983)a
                                                                               
    Source category                                  Emission rate (tonnes/year)
                                                                               

    Coal combustion
    - electric utilities                                          780-4650
    - industry and domestic                                       990-9900

    Oil combustion
    - electric utilities                                          230-1740
    - industry and domestic                                       720-2150

    Pyrometallurgical non-ferrous metal production
    - mining                                                     1700-3400
    - lead production                                        11 700-31 200
    - copper-nickel production                               11 000-22 100
    - zinc-cadmium production                                  5520-11 500

    Secondary non-ferrous metal production                         90-1440

    Steel and iron manufacturing                               1070-14 200

    Refuse incineration
    - municipal                                                  1400-2800
    - sewage sludge                                                240-300

    Phosphate fertilizers                                           60-270

    Cement production                                            20-14 200

    Wood combustion                                              1200-3000

    Mobile sourcesb                                                248 030

    Miscellaneous                                                3900-5100

    Total                                                  289 000-376 000

                                                          (median 332 000)

                                                                               

    a  Adapted from: Nriagu & Pacyna (1988), as in OECD (1993).
    b  Current estimates (OECD, 1993) for mobile source emissions
       would be about 30% of the 1983 amounts.

    Table 8.  Worldwide emissions of lead into soils (1983)
                                                                               
    Source category                                  Emission rate (tonnes/year)
                                                                               

    Agricultural and food wastes                               1500-27 000
    Animal wastes, manure                                      3200-20 000
    Logging and other wood wastes                                6600-8200
    Urban refuse                                             18 000-62 000
    Municipal sewage sludge                                      2800-9700
    Miscellaneous organic wastes, including excreta                20-1600
    Solid wastes, metal manufacturing                          4100-11 000
    Coal fly ash, bottom fly ash                            45 000-242 000
    Fertilizer                                                    420-2300
    Peat (agricultural and fuel use)                              450-2000
    Wastage of commercial products                         195 000-390 000
    Atmospheric fall-out                                   202 000-263 000
                                                                          

    Total yearly input to soils                          479 090-1 038 800

    Mine tailings                                          130 000-390 000

    Smelter slags and wastes                               194 000-390 000
                                                                          

    Total yearly discharge on land                       803 090-1 818 800
                                                                               

    a  From: Nriagu & Pacyna (1988), adapted from OECD (1993);
       many of these emissions remain localized due to the nature of
       the particulate matter
    
    4.  ENVIRONMENTAL TRANSPORT, DISTRIBUTION AND TRANSFORMATION

         Over the last 10-15 years, a great many studies have been
    conducted on the complex interrelationships between environmental lead
    emissions and their deposition on such environmental surfaces as
    vegetation, soil, house dust and water. All are potential sources of
    lead exposure for humans. Transport between environmental compartments
    also takes place (see Fig. 2). A full discussion of the complex
    physical and chemical processes controlling these pathways is beyond
    the scope of this monograph and the reader is directed towards other
    reviews for more details (Elias, 1985; US EPA, 1986a; IPCS, 1989). 

    4.1  Transport and distribution between media

    4.1.1  Atmospheric deposition

         From the mass balance point of view the transport and
    distribution of lead from major emission sources, both fixed and
    mobile, is mainly atmospheric. Most of the lead discharged to the
    atmosphere is deposited near the source. However, approximately 20% is
    widely dispersed (Nriagu, 1979; IPCS, 1989) and contaminates areas as
    remote as glacial strata in Greenland (Settle & Patterson, 1980). The
    extent of long-range transport of lead particles is dependent upon
    particle size, particles > 2 µm in diameter being deposited close to
    the source of emission. Between 20 and 60% of emissions from vehicles
    has been reported to remain within 25 m of the roadway (ATSDR, 1991).
    However, in view of the marked decrease in the concentration of lead
    in cores of ice from Greenland since the decreased use of leaded
    petrol (Boutron et al., 1991), it is apparent that vehicle emissions
    can contribute to the levels of lead in air far from the source.
    Long-range transport of lead particles was also noted by Evans &
    Rigler (1985).

         Lead can be removed from the atmosphere and transferred to
    environmental surfaces and compartments by wet or dry deposition. Wet
    deposition appears to be more important than dry deposition for the
    removal of atmospheric lead. Depending upon geographical location and
    the level of emissions in the area, between 40 and 70% of atmospheric
    lead is removed by wet deposition (Nielsen, 1984). In most cases it is
    poorly soluble and either precipitates out in soils and sediments or
    is bound to organic matter in these compartments. For these reasons
    lead is not readily removed and tends to accumulate in those
    ecosystems where it is deposited (IPCS, 1989). Chan et al. (1986)
    calculated the ratio of wet to dry deposition to be 1.63, 1.99 and
    2.50 for sites in south, central, and northern Canada, respectively,
    while Talbot & Andren (1983) reported that wet deposition accounted
    for 80% of the total lead deposited in a semi-remote site in the USA.

    FIGURE 2

         Making several assumptions regarding global atmospheric lead
    concentrations, wind speed, surface area and texture, a global
    deposition of approximately 410 000 tonnes/year (combined wet and dry)
    was calculated by the US EPA (1986a).

    4.1.2  Transport to water and soil

         When deposited in water, whether from air or through run-off from
    soil, lead partitions rapidly between the sediment and aqueous phase,
    depending upon the salt content of the water as well as the presence
    of organic complexing agents. For example, at pH > 5.4 the total
    solubility of lead is about 30 µg/litre in hard water and 500 µg/litre
    in soft water (Davies & Everhart 1973). In addition, the presence of
    sulfate and carbonate ions can limit lead solubility, as described by
    Hem & Durum (1973) in a review of the aqueous chemistry of lead.

         Water-borne lead has been found to exist as soluble lead or
    undissolved colloidal particles, either suspended in the aqueous phase
    or carried as surface coatings on other suspended solids (Lovering,
    1976). The ratio of lead in suspended solids to lead in the dissolved
    form has been found to vary from 4:1 in rural areas to 27:1 in urban
    streams (Getz et al., 1977).

         Both natural organic compounds (humic and fulvic acids) as well
    as those of anthropogenic origin (e.g., ethylenediamino-tetraacetic
    and nitrilotriacetic acids) may complex lead found in surface waters
    (Steelink, 1977; Reuter & Perdue, 1977; Neubecker & Allen, 1983). The
    presence in water of such chelators can increase the rate of solution
    of lead compounds (e.g., lead sulfide) 10 to 60 times over that of
    water at the same pH without fulvates (Bondarenko, 1968; Lovering,
    1976).

         Lead accumulation in soils is primarily a function of the rate of
    wet and dry deposition from the atmosphere. Transport within soil and
    the bioavailability of lead from soil are dependent upon many factors,
    including pH, mineral composition of the soil, and amount and type of
    organic material, with most of the lead being bound within the upper
    5 cm of soil (Reaves & Berrow 1984; Garcia-Miragaya, 1984). This
    limits the amount which can be leached into water or be available for
    uptake into plants. It has been shown that only 0.2% of the total lead
    in soil can be released into solution by shaking (Dong et al., 1985).
    However, the release of lead from organic complexes to the soluble,
    and thus bioavailable, form is highly pH dependent. Within the usual
    pH range for soils (4 to 6), the organic-lead complexes become more
    soluble and the lead more available for plant uptake and leaching into
    water (US EPA, 1986a). 

    4.1.3  Transport to biota

         The transfer of air lead to the biota may be direct or indirect
    (uptake from water, soil and vegetation). Examples of the accumulation

    of lead into aquatic (wet and dry deposition) and terrestrial
    organisms are given in Environmental Health Criteria 85: Lead -
    Environmental Aspects (IPCS, 1989). Relevant parts of that monograph
    are summarized here. 

    4.1.3.1  Aquatic organisms

         In aquatic and aquatic/terrestrial model ecosystems, uptake by
    primary producers and consumers seems to be determined by the
    bioavailability of the lead. Bioavailability is generally much lower
    whenever organic material, sediment or mineral particles (e.g., clay)
    are present. In many organisms, it is unclear whether lead is adsorbed
    onto the organism or actually taken up. Consumers take up lead from
    their contaminated food, often to high concentrations but without
    biomagnification.

         The uptake and accumulation of lead by aquatic organisms from
    water and sediment are influenced by various environmental factors
    such as temperature, salinity and pH, as well as humic and alginic
    acid content.

         In contaminated aquatic systems, almost all of the lead is
    tightly bound to sediment. Only a minor fraction is dissolved in the
    water, even in the interstitial water.

         The lead uptake by fish reaches equilibrium only after a number
    of weeks of exposure. Lead is accumulated mostly in gill, liver,
    kidney and bone.

         Fish eggs show increasing lead levels with increased exposure
    concentration, and there are indications that lead is present on the
    egg surface but not accumulated in the embryo.

         Fish accumulate lead from water as well as sediments; aquatic
    uptake is influenced by the presence of cations and the oxygen content
    of the water (IPCS, 1989).

    4.1.3.2  Terrestrial organisms

         In bacteria, the majority of lead is found in the cell wall. A
    similar phenomenon is also noted in higher plants. Some lead that
    passes into the plant root cell can be combined with new cell wall
    material and subsequently removed from the cytoplasm to the cell wall.
    Of the lead remaining in the root cell, there is evidence of very
    little translocation to other parts of the plant because the
    concentration of lead in shoot and leaf tissue is usually much lower
    than in root. Foliar uptake of lead occurs, but only to a very limited
    extent.

         In animals, there is a positive correlation between tissue and
    dietary lead concentrations, although tissue concentrations are almost
    always lower. The distribution of lead within animals is closely
    associated with calcium metabolism.

         Lead shot is typically trapped in the gizzard of birds where it
    is slowly ground down resulting in the release of lead.

         The tetravalent organic form of lead is generally more toxic than
    the divalent inorganic form, and its distribution in organisms may not
    specifically follow calcium metabolism.

    4.2  Environmental transformation

    4.2.1  Abiotic transformation

         Once released into the environment lead may be transformed from
    one inorganic species or particle size to another. However, as an
    element it is not subject to degradation. For example, lead-containing
    particles in automobile exhaust are usually lead halides or double
    salts with ammonium halides (Biggins & Harrison, 1979). Within 24 h,
    over 75% of lead particulate matter is transformed to lead carbonates
    and sulfates (Olson & Skogerboe, 1975).

    4.2.2  Biotransformation

         The transformation of inorganic lead to tetramethyllead (TML) has
    been observed in aquatic systems, particularly in sediments, and
    biomethylation was postulated by Wong et al. (1975) and Schmidt &
    Huber (1976). However, no biological methylation of inorganic lead was
    noted by Reisinger et al. (1981) in studies under many conditions
    using several bacterial species known to alkylate mercury and other
    heavy metals. The authors did find chemical methylation in the
    presence of methylcobalamin and sulfide or aluminium ions and it was
    independent of the presence of bacteria. The evidence for microbial
    methylation of various compounds of lead in aquatic systems has been
    reviewed by Beijer & Jernelöv (1984). It is still unclear whether the
    TML formed is produced abiotically or by biotransformation.

    5.  ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE

         This chapter describes the sources of lead to which people are
    exposed and quantifies that exposure. Depending on the source, the
    concentration of lead and its bioavailability, the relative
    contribution of each source may vary considerably. For example, men
    working in ship-breaking, children exposed to deteriorating lead
    paint, and people consuming soft water distributed in lead pipes have
    frequently been shown to have excessive absorption of lead, leading to
    clinically obvious lead poisoning.

         In evaluating the exposure of the general population to lead, it
    is critical to consider the interrelationships among environmental
    pathways for lead and transfers across environmental media. The
    general population is exposed to lead simultaneously from many sources
    and through multiple pathways as shown in Fig. 2. Thus, while, for
    purposes of discussion, exposure via air, water, food, dusts and
    soils, and other sources are presented separately in this document,
    the total exposure of the general population from all sources must be
    considered. Those working in industries where lead is used or produced
    may be subject to additional exposure compared with the general
    population.

         In addition, exposure of certain groups within the general
    population may vary because of physiological, behavioural or other
    factors. For example, the fetus is exposed to lead via maternal
    transfer of internal and external doses, nursing infants may be
    exposed to lead in breast milk, the young child is exposed more
    intensively to dusts and lead on non-food items (such as lead-painted
    toys), alcohol consumption and smoking increase lead exposure,
    differences in diet may influence lead exposure markedly, and some
    people may be exposed to lead through hobby or occupational activities
    in addition to their exposure as members of the general population.

    5.1  Inhalation route of exposure

    5.1.1  Ambient air

         Ambient air can be a major pathway of lead distribution in the
    environment. Sources of lead in air include combustion products of
    lead additives in petrol, and point sources such as smelters,
    incinerators, and some industrial processes including the burning of
    fossil fuels.

         Concentrations of lead in air range from 7.6 × 10-5 µg/m3 in
    remote areas such as Antarctica (Maenhaut et al., 1979) to
    > 10 µg/m3 near lead smelters (Elias, 1985).

         Almost all lead in air is bound to fine particles of less than
    1 µm diameter, although some may be solubilized in acid aerosol
    droplets. The size of these particles varies with the source and with

    the age of the particle from the time of emission (US EPA, 1986a; WHO,
    1987). Most lead in air is inorganic lead, and the predominant source
    is from the combustion of tetraethyl- and tetramethyllead used as fuel
    additives (US EPA, 1986a; WHO, 1987). A summary of lead levels on fine
    airborne particles from some cities in the USA and France is given in
    Table 9.


    Table 9.  Concentrations of lead in fine airborne particles from some
              cities in the USA and France in 1984-1985a
                                                                        
    City                    Population        Mean lead          No. of
                                              concentration      samples
                                              (µg/m3)          
                                                                        

    Clemson (USA)               3000             0.33               15
    Senonches (France)          3000             0.005               6
    Orleans (France)         110 000             0.11                7
    Clermont (France)        161 000             0.045               7
    Akron (USA)              200 000             0.052               6
    Strasbourg (France)      260 000             0.072               7
    Norfolk (USA)            270 000             0.031               6
    Chicago (USA)         10 000 000             0.064               5
    Paris (France)        10 000 000             0.44                7
                                                                        

    a  From: Delumyea & Kalivretenos (1987)


    5.1.1.1  Emissions from motor vehicles

         In Europe, where leaded vehicle fuel is still used, airborne
    concentrations of lead in urban areas are likely to be in the range of
    0.5-3 µg/m3 (WHO, 1987). Concentrations of between 0.6 and
    5.7 µg/m3 were reported in Mexico in 1982 (GEMS, 1985). Where leaded
    vehicle fuel is no longer used, concentrations are likely to fall to
    < 0.2 µg/m3 (Elias, 1985). In 1990, concentrations of lead in air
    in urban areas of the USA had fallen to below 0.07 µg/m3 (US EPA,
    1991). This decrease in airborne lead levels in the USA is shown in
    Fig. 3. Reductions in lead in air have been reported from Canada,
    Germany, Norway and the United Kingdom (OECD, 1993).

    5.1.1.2  Stationary sources

         Where emissions are largely uncontrolled, concentrations of lead
    in air around stationary sources such as lead smelters range from over
    10 µg/m3 50 m from the smelter to 1.5 µg/m3 one km away (Wang et
    al., 1992). Where more stringent emission controls are used,
    concentrations of lead are much lower (US EPA, 1991). The Port Pirie

    FIGURE 3

    smelter in South Australia is estimated to have lost 80 000 tonnes of
    lead to the environment in non-stack fugitive emissions from 1889 to
    1982. It is also estimated that from 1969 to 1981 this smelter
    discharged 40 tonnes of lead per year into the environment. Ambient
    air concentrations near the smelter were between 0.5 and 10 µg/m3
    (Body et al., 1988).

    5.1.2  Indoor air

         Davies et al. (1987a) sampled indoor and ambient air lead levels
    and found that where there was no interior lead source, such as
    lead-painted surfaces, air lead concentrations inside dwellings were
    highly correlated with those outside and averaged approximately 60% of
    those in the external air immediately outside the house. A similar
    ratio was reported in the Arnhem lead study (Diemel et al., 1981).
    Indoor air lead levels are affected by the presence of smokers and
    lead-painted surfaces.

         Levels of airborne lead in indoor shooting ranges have been shown
    to range from 2.7 to 90.5 µg/m3 depending upon the location of
    sampling, which varied from the showroom to target area (Novotny et
    al., 1987).

    5.1.3  Air in the working environment

         The diversity and extent of the industrial applications of lead
    is such that it is impossible to make general statements about
    exposure levels. In many instances actual exposure levels have not
    been measured and often work is carried out in small enterprises which
    may not be subject to workplace controls or legislated requirements.

         Airborne lead concentrations in the occupational setting vary
    considerably according to the type of industry and the level of
    industrial hygiene practised at each plant. Occupations and operations
    that may present lead hazards to workers are listed in Table 10.
    Recent monitoring data (1980-1985) from Finland are summarized in
    Fig. 4 (Jaakkola & Anttila, 1992).

    5.1.4  Smoking of tobacco

         Lead is present in tobacco. The mean content of lead in
    filter-tipped cigarettes produced between 1960 and 1980 was 2.4 µg/g.
    Approximately 5% of this lead may be inhaled; the remainder occurs in
    the ash and side-stream smoke (Mussalo-Rauhamaa et al., 1986).

    FIGURE 4

        Table 10.  Occupations or operations which may present lead hazards for
               workersa
                                                                               

    Primary and secondary lead smelting                   Lead mining

    Welding and cutting of lead-painted metal             Plumbing
    constructions

    Welding of galvanized or zinc silicate coated sheets  Cable making

    Shipbreaking                                          Wire patenting

    Nonferrous foundries                                  Lead casting

    Storage battery manufacture: pasting, assembling,     Type founding in
    welding of battery connectors                         printing shops

    Production of lead paints                             Stereotype setting

    Spray painting                                        Assembling of cars

    Mixing (by hand) of lead stabilizers into             Automobile repair
    polyvinyl chloride

    Mixing (by hand) of crystal glass mass                Shot making

    Sanding or scraping of lead paint                     Welding (occasionally)

    Burning of lead in enamelling workshops               Lead glass blowing

    Repair of automobile radiators                        Pottery/glass making
                                                                               

    a  Adapted from: Hernberg (1973)
    
    5.2  Exposure by ingestion

    5.2.1  Water

         Exposure of humans to lead from water has been underestimated in
    studies of total exposure. Due to the practice of sampling water
    systems at points before entry into the distribution piping and
    domestic plumbing (US EPA, 1986a; Dabeka et al., 1987), it had been
    widely assumed that exposure to lead in drinking-water was not
    significant.

         Background or natural levels of lead in surface and ground water
    are generally low. However, water with low pH and only low
    concentrations of dissolved salts (referred to as aggressive) can
    leach substantial quantities of lead from pipes, solder and fixtures.
    Lead-lined reservoirs, cisterns and holding tanks for water (Mushak &
    Crocetti, 1989) can be a major source of lead contamination of
    drinking-water. For example, Wiebe et al. (1991) reported the analysis
    of over 2000 water samples in Hawaii, USA, following increased
    volcanic activity that resulted in the release of acid aerosols. The
    lead concentration of drinking-water collected in catchment systems
    ranged from < 20 to 7000 µg/litre. Sampling programmes conducted at
    the tap in the USA during 1985-1988 revealed widespread elevation of
    lead in drinking-water, often above the WHO guidance value of
    50 µg/litre (WHO, 1984), which has now been revised to 10 µg/litre
    (WHO, 1993).

         The combination of acid or aggressive water and lead in plumbing
    results in very high concentrations of lead in drinking-water,
    particularly after it has been standing for several hours (Worth et
    al., 1981; Sherlock et al., 1982; Kaminsky et al., 1988).

         Surveys in Canada and the USA showed that drinking-water supplies
    leaving treatment plants contain 2-3 µg lead/litre (US EPA, 1986a;
    Dabeka et al., 1987). In the case of plumbosolvent water, up to 40% of
    household samples may exceed 100 µg lead/litre. This has been observed
    in Scotland (Sherlock et al., 1986), and reflects the contribution of
    plumbing and plumbing fixtures to lead levels of drinking-water.

    5.2.2  Food and alcoholic beverages

    5.2.2.1  Food

         The major source of lead for non-occupationally exposed adults is
    food and drink. The proportion of total intake derived from food is
    dependent on the concentration of lead in air, water and other
    sources. Detailed data are available from several countries, including
    Australia (NFA, 1991), USA (Bolger et al., 1991), Sweden (Vahter et
    al., 1990) and Canada (Dabeka et al., 1987). Foods have been surveyed
    from several other industrialized and developing countries
    (Galal-Gorchev, 1991b). Children are exposed to additional lead from
    dust and soil, and so lead from foods and beverages may not be the
    predominant sources of lead for all age groups.

         Lead is present in soils and is transferred to food crops growing
    on the soil. Roots usually contain more lead than stems and leaves,
    while seeds and fruits have the lowest concentrations.

         Particulate lead present in air may adhere tenaciously to leafy
    vegetables. Leaves collected in or very near urban areas have been
    shown to contain substantially elevated concentrations of lead.
    Quantities of lead ingested from the diet vary widely from country to
    country.

         Data on the lead levels of specific foodstuffs or groups of food
    materials, from which one can estimate a daily dietary lead intake,
    are available from several countries. In a few studies, foodstuffs
    specific to infants and children have also been analysed (Kolbye et
    al., 1974; Dabeka & McKenzie, 1987; Vahter et al., 1990; Bolger et
    al., 1991; Albert & Badillo, 1991). Data are available for canned
    foods typically consumed by young children (Capar & Rigsby, 1989). The
    utilization of such data in the calculation of total intake of lead
    from food is discussed in section 5.2.2.2.

         An overview of the foods contributing to the dietary lead levels
    in Australia is shown in Fig. 5. Similar data from other countries are
    shown in Table 11 (Galal-Gorchev, 1991b). A summary of the data on
    lead levels in foodstuffs from the USA is given in Table 12 and from
    Canada in Table 13. More recent data from the USA have shown that
    there has been a substantial reduction of lead levels in food consumed
    by all age groups during the past two decades (Bolger et al., 1991)
    (Fig. 6). A similar decrease in lead intake has been found in the
    United Kingdom (OECD, 1993). These data should be considered as
    representative of specific areas and such values can be expected to
    vary elsewhere according to local agricultural and food-processing
    practices, particularly in areas where lead-soldered cans are still
    used. Support for this is indicated by a comparison on the one hand of
    the decreased use of lead-soldered food and beverage cans in the USA
    (Fig. 7) and the increased use of cans produced by alternative
    technology, and on the other hand the marked decrease in food-borne
    lead shown in Fig. 6. Food may represent a pathway for human lead
    exposure from other media such as air and water. The use of leaded
    gasoline or the proximity of industries that may produce ambient
    emissions of lead can greatly influence dietary lead intake. Therefore
    further caution is required when extrapolating between countries with
    regards to levels of food-borne lead.

         Representative levels of lead in foodstuffs from some 20
    countries are given in Table 14 (Galal-Gorchev, 1991a). These results
    from the GEMS/FOOD data can be compared with the levels of lead in
    specific foodstuffs in the USA and Canada (Tables 12 and 13).

    FIGURE 5

    FIGURE 6

    FIGURE 7

    Table 11.  Foods contributing to dietary lead levels in Canada,
               Finland, Netherlands and the United Kingdoma
                                                                        
    Country          Food                    Percentage of total intake
                                                                        

    Canada           vegetables                        17
                     meat/fish/poultry                 17
                     beverages                         15
                     cereals and products              15
                     fruits and juices                 10

    Finland          cereals and products              24
                     fruits                            22
                     beverages, sweets, etc.           20
                     milk and products                 17
                     vegetables                         9

    Netherlands      drinking-water                    30
                     cereals and products              17
                     vegetables                        12
                     wines and spirits                  9
                     fruits                             6

    United Kingdom   bread and cereals                 15
                     beverages                         14
                     potatoes                          10
                     milk                               9
                     canned vegetables                  8
                                                                        

    a  From: Galal-Gorchev (1991b)

    Table 12.  Concentrations of lead in various foods in the USAa
                                                                   

    Food group                         Concentration of leadb (µg/g)
                                                                   
    Dairy products                              0.003-0.083
    Meat, fish and poultry                      0.002-0.159
    Grain and cereal products                   0.002-0.136
    Vegetables                                  0.005-0.649
    Fruit and fruit juices                      0.005-0.223
    Oils, fats and shortenings                  0.006-0.073
    Sugar and adjuncts                          0.006-0.073
    Beverages                                   0.002-0.041
                                                                   

    a  From: US EPA (1986a)
    b  Range of concentrations shown are the lowest and highest mean
       values for items within the food group and listed in Appendix 7-D
       of US EPA (1986a).

        Table 13.  Levels of lead in various Canadian food categoriesa
                                                                                     
         Description of food category                             Median concentration
                                                                  of lead (µg/kg)b
                                                                                     

    I     Cereals (as prepared with milk, sugar, etc.),           32.4 (11.5-78.3)
          bread and toast

    II    Water consumed directly                                 2.0 (0.25-71.2)

    III   Coffee, tea, beer, liquor, sodas, etc. (as prepared)    8.8 (< 0.05-28.9)

    IV    Fruit juices, fruits (canned and fresh)                 7.9 (1.5-109)

    V     Dairy products and eggs                                 3.3 (1.21-81.9)

    VI    Starch vegetables, e.g. potatoes, rice                  16.9 (5.5-83.7)

    VII   Other vegetables, vegetable juices and soups            31.7 (0.62-254)

    VIII  Meat, fish, poultry, meat-based soups                   31.3 (11-121)

    IX    Miscellaneous (pies, puddings, nuts, snack foods)       33.1 (13.6-1381)

    X     Cheese (other than cottage cheese)                      33.8 (27.7-6775)
                                                                                     

    a  From: Dabeka et al. (1987)
    b  Values in parentheses are the ranges.
    

         The lead levels in infant foods in Canada, Mexico and USA are
    shown in Table 15. In 1987, Dabeka et al. (1987) found the intake of
    lead by infants fed evaporated milk stored in lead-soldered cans
    exceeded the Provisional Tolerable Weekly Intake of 25 µg lead/kg body
    weight, set in 1993 (FAO/WHO, 1993). These values do not include lead
    in water used to prepare formulae. It has been reported that infants
    fed formulae prepared with water containing high levels of lead
    (> 100 µg/litre) have lead intakes exceeding 25 µg/kg body weight per
    week (Galal-Gorchev, 1991b).

    5.2.2.2  Total intake from food

         Guidelines for the determination of dietary intake of chemical
    contaminants have been published (WHO, 1985). Three basic approaches
    were described, namely: (i) total diet (i.e. market or shopping
    basket) studies; (ii) selective studies of individual foodstuffs, and

    (iii) duplicate portion (duplicate diet) studies. It is essential to
    have food consumption data for the first two methods in order to
    estimate a total intake. For all methods, well-designed quality
    assurance and quality control programmes are essential and these have
    been described by Vahter & Slorach (1990) and Vahter et al. (1991a).


    Table 14.  Representative levels of lead in foods from GEMS/FOOD
               dataa
                                                                      
    Commodity                             Typical lead levels (µg/kg)
                                                                      

    Cereals                                           60
    Roots and tubers                                  50
    Fruit                                             50
    Vegetables                                        50
    Meat                                              50
    Vegetable oils and fats                           20
    Fish                                             100
    Pulses                                            40
    Eggs                                              20
    Nuts and oilseeds                                 40
    Shellfish                                         20
    Offal                                             20
    Spices and herbs                                  30
    Other foods                                  not assessed
    Drinking-water                                    20
    Canned beverages                                 200
    Canned foodb                                     200
                                                                      

    a  From: Galal-Gorchev (1991a)
    b  It is assumed that canned food consumption is 2% of total.


         Given the differences between countries with respect to dietary
    composition, the amount of specific foodstuffs consumed, the
    processing technologies employed, whether consumption of water and
    alcoholic beverages are included in the estimates of dietary lead, and
    the number of samples taken, caution must be exercised in making
    comparisons between countries.

         During the late 1970s and 1980s the quantity of lead ingested as
    part of diet decreased markedly in many countries. For adults not
    occupationally exposed to lead, the diet remains the largest
    contributor to lead intake. However, the quantities ingested are far
    lower than in previous decades. In the USA, typical levels of intake
    declined as shown in Fig. 6.

        Table 15.  Lead levels (µg/kg food) in cow's milk and infant formula
                                                                             

    Product                            Canada          Mexico        USA
                                    median (range)a    averageb    averagec
                                                                             

    Fluid milk                      1.19 (0.01-2.5)       5

    Evaporated milk (canned)        71.9 (27-106)        88          10
    (cardboard)                          --               9

    Infant formula
     Ready to use lead-solder can   30.1 (1.1-122)       13          10
     Ready to use lead-free can     1.6 (1.5-2)           1

    Formula powder (1985)           96.6 (3.7-19)

    Powdered milkd                       --              21
                                                                             

    a  From: Dabeka & McKenzie (1987)
    b  From: Albert & Badillo (1991). Data were obtained in 1982.
    c  From: Bolger et al. (1991). Data were obtained in the late 1980s.
    d  The concentration of lead in milk consumed by the infant will be
       highly dependent on the concentration of lead in water used to
       dilute the powdered milk.
    
         An overall summary of the GEMS/FOOD data for adults is given in
    Fig. 8 (Galal-Gorchev, 1991a). The trends for lead intake for the
    period 1980-1988 in the USA, Japan, Hungary and the United Kingdom are
    shown in Fig. 9 (Galal-Gorchev, 1991b). Data for intake of lead by
    infants and children in eight countries are shown in Fig. 10
    (Galal-Gorchev, 1991a). Additional data are available from other
    countries. Although collection methods vary, these data illustrate the
    wide variations in ingestion of lead from food. Brunekreef (1986)
    noted that the market basket studies tended to overestimate lead
    intake when compared with duplicate diet analysis. Examples cited
    included reports from the United Kingdom by Fouassin & Fondu (1980),
    Buchet et al. (1983) and Sherlock et al. (1982), where market basket
    surveys overestimated this intake by 2- to 3-fold.

         Other studies from various countries on total lead intakes by
    children and adults are summarized in Table 16. Lead contaminated
    water has been shown to be a contributor to food-borne lead where
    large volumes of water are used in food preparation and cooking, e.g.,
    in foods prepared in boiling water.

    FIGURE 8

    FIGURE 9

    FIGURE 10

        Table 16.  Daily lead intake via food in adults and children
                                                                                      
    Population studied            Daily intake (µg/day)a      Reference
                                                                                      

    Adults, United Kingdom             110    M               Brunekreef (1986)
    Adults, United Kingdom              71    D               Brunekreef (1986)
    Adults, Belgiumb                   282    M               Fouassin & Fondu (1980)
    Adults, Belgiumb                    96    D               Buchet et al. (1983)
    Adults, Sweden                      27    M               Slorach et al. (1983)
    Adults, Finland                     66    M               Varo & Kovistoinen (1983)
    Adults, Canada                      43    D               Dabeka et al. (1987)
    Adults, USA                         82    M               Gartrell et al. (1985a)
    Adults (female), Japan              31    D               Vahter et al. (1991b)
    Adults, Germany                     61                    Kampe (1983)
    Adults (female), Croatia            15    D               Vahter et al. (1991b)
    Adults, Italy                      140                    IAEA (1987)
    Adults (female), China              46    D               Vahter et al. (1991b)
    Adults, Turkey                      70                    IAEA (1987)
    Adults (female), Sweden             26    D               Vahter et al. (1991b)
    Children, Poland
     0-1 year                          225                    Olejnik et al. (1985)
     1-3 years                         259
     7-18 years                        316
    Adults, New Zealand                316    M               Pickston et al. (1985)
    Children (infant), UK              2-3    breast milk     Kovar et al. (1984)
    Children (< 1 year), USA         16-17    infant formula  Ryu et al. (1983)
                                              & milk
    Children, USA
     6 months                         33.5    M               Gartrell et al. (1985b)
     2 years                          43.4
                                                                                      

    a  M = Market basket survey; D = Duplicate diet study
    b  Populations studied from the same region.
    

         The relative intake of lead from various sources in 1986 and 1990
    in 2-year-old infants and women of child-bearing age in the USA is
    shown in Fig. 11. These data illustrate the marked decrease in lead
    intake from food over a 4-year period in which there were marked
    reductions in the use of lead-soldered cans and lead-containing petrol
    additives in the USA (Bolger et al., 1991). Similar decreases in other
    countries would no doubt occur after similar actions by public health
    officials.

    FIGURE 11

    5.2.2.3  Alcoholic beverages

         Contamination of alcoholic beverages with lead may occur in
    several way. For example, lead solder used to repair casks or keg and
    tap lines from lead capsules used as seals or from residues of lead
    arsenate pesticides in soils now used to grow grapes. Alcoholic
    beverages tend to be acidic and there is the possibility that large
    amounts of lead can be dissolved during preparation, storage or
    serving (Wai et al., 1979). Published reports on lead levels in wine
    show that important variations occur from sample to sample (Jorhem et
    al., 1988). The US Department of the Treasury (1991) analysed 432
    table wines sold within the USA. The results are summarized in Table
    17.

         Sherlock et al. (1986) found that the majority of canned and
    bottled beer (90 and 86% respectively) contained less than 10 µg
    lead/litre. Draught beers typically contained higher lead
    concentrations with 55% having lead concentrations greater than
    10 µg/litre, 16% with concentrations over 20 µg/litre, and 4% with
    concentrations over 100 µg/litre. The higher lead concentrations in
    draught beers are considered most likely due to the draught-dispensing
    equipment which sometimes contains brass or gunmetal, both of which
    contain low but significant amounts of lead (Sherlock et al., 1986).

         In general, alcoholic beverages do not appear to be a significant
    source of lead intake for the average person.

    Table 17.  Distribution of lead in table wines in USAa
                                                         
    Range          Number of           Percentage of total
    (µg/litre)     samples             samples analysedb
                                                         

       0-10           36                     8.3
      11-25           62                    14.4
      26-50          105                    24.3
     51-100          144                    33.3
    101-250           64                    14.8
    251-500           12                     2.8
    501-673            9                     2.1
                                                         

    a  From: US Department of the Treasury (1991)
    b  In all, 432 samples were analysed.

    5.2.3  Dust and surface soils

    5.2.3.1  Dust

         Dust is a significant source of exposure to lead, particularly
    for young children (see Fig. 11), as has been demonstrated in several
    studies correlating children's blood lead concentrations with dust
    lead levels (Rabinowitz et al., 1985; Bornschein et al., 1987; Davies
    et al., 1987a; Laxen et al., 1987; Steenhout, 1987).

         The major contributions to lead levels in soil and outdoor dust
    are from the combustion of fossil fuels (principally leaded petrol),
    stationary sources such as smelters, and peeling and flaking of
    lead-based paint. Typical lead levels in road dust in the USA are
    800-1300 mg/kg in rural areas to 100-5000 mg/kg in urban areas (US
    EPA, 1989c).

         Concentrations of lead in household dust vary greatly between
    different dwellings and areas of the world. Mean concentrations of
    300-2500 mg/kg have been found in the United Kingdom and USA, but
    individual samples may be in the range of 10 000 to 30 000 mg/kg (Que
    Hee et al., 1985b; Clark et al., 1985; Bornschein et al., 1986; Raab
    et al., 1987).

         Flaking lead-based paint, paint chips, and weathered powdered
    paint markedly increase intake of lead from surface dust, particularly
    for urban children with pica (US EPA, 1986a; Bornschein et al., 1986).
    Lead-based paint chips have been found to contain 1000-5000 µg
    lead/cm2 (Billick & Gray, 1978). When lead-based paint is present,
    interior renovation activities greatly increase household dust lead
    concentrations (Laxen et al., 1987). Improved control of dust and
    surface clean-up after lead-based paint removal have been shown to
    reduce lead exposure of children reoccupying affected houses (Charney
    et al., 1983).

    5.2.3.2  Soil

         In rural and remote areas, lead in soil is derived mainly from
    natural geological sources. These natural sources account for 1-30 mg
    lead/kg, but where soils are derived from leadmineralized rocks,
    natural concentrations may range from several hundred to several
    thousand mg/kg.

         Typical values for lead in rural soils in the United Kingdom are
    15-106 mg/kg with a geometric mean of 42 mg/kg (Davies, 1983). A
    geometric mean of 48 mg/kg for 2780 samples has also been reported
    (McGrath, 1986).

         Concentrations of lead in urban soil vary greatly. In the USA, a
    study of city parks recorded concentrations of 200 to 3300 mg/kg (US
    EPA, 1989). Concentrations of up to 10 960 mg/kg have been reported

    for urban garden soils in the USA (Mielke et al., 1984), and up to 14
    100 mg/kg in the United Kingdom (Culbard et al., 1988). Concentrations
    can exceed 20 000 mg/kg around lead mining and processing operations
    (Culbard et al., 1988). In areas where lead-based paint has been used,
    soil samples taken near building foundations have been reported to be
    as high as 20 000 mg/kg (Schmitt et al., 1988; Krueger & Duguay,
    1989).

         In general, lead concentrations in soils near roads are high
    where road traffic density is high. Concentrations decrease
    exponentially with distance from the road (IPCS, 1989).

         Continuous application of sewage sludge results in an
    accumulation of lead in soil. For example, soil receiving heavy
    applications over a long period was found to contain 425 mg/kg,
    compared with 47 mg/kg in an untreated soil (Beckett et al., 1979).

    5.2.3.3  Migration of lead from food containers

         The available data on the daily intake of lead by adults and
    children indicate a general decrease in those areas where the level of
    lead in petrol has decreased and a concerted effort made to avoid
    lead-soldered cans for food storage (OECD, 1993). However, in many
    regions of the world, lead can migrate from food storage and serving
    vessels such as lead-soldered cans (see section 5.2.2.2), ceramic
    dishes, pottery vessels, crystal glassware and decals on food wrap
    and/or dishes. Acidic foods tend to leach more lead. However, certain
    foods such as corn and beans are associated with greater release of
    lead than would be predicted from their acidity alone (Bolger et al.,
    1991). Oxygen appears to accelerate the release of lead from food
    containers.

         If foods are stored in ceramic or pottery dinnerware that was
    lead-glazed and fired in a low temperature kiln, lead can migrate from
    the pottery glaze into the food. The glazing process uses a flux,
    which is a material that, at high temperatures, reacts with and helps
    dissolve the components of the glaze. Lead oxide is a commonly used
    flux. Factors that determine whether and to what extent lead will
    migrate include the temperature and extent of firing of the pottery in
    the manufacturing process, temperature and duration of food storage,
    and the acidity of the food. It is extremely difficult to quantify the
    extent of such exposures in view of the variations in the
    manufacturing processes and the quality control practised in the
    country of origin. However, the extent of exposure can be quite
    significant, particularly among infants. Cases of lead intoxication
    from this source have been reported (Wallace et al., 1985).

         Lead has been found to migrate from lead crystal glass into
    beverages. This problem is especially severe if beverages are stored
    in lead-crystal containers, e.g., decanters or liquor bottles (de
    Leacy, 1987; Graziano & Blum, 1991). This phenomenon was not observed
    with borosilicate glass containers (de Leacy, 1987).

         Several studies have been made of lead contamination of foods and
    beverages from lead used in the manufacture or repair of metal
    vessels. Recoating the inner surface of brass utensils with a mixture
    of lead and tin, described as "tinning", is widely practised by
    artisans in India (Vatsala & Ramakrishna, 1985). The tin-lead alloy
    contains 55 to 70% lead. Water containing tamarind contained
    400-500 µg lead/litre after 5 min of boiling. The practice is
    considered to be widely prevalent in at least the three southern
    states of India. Zhu (1984) described 344 cases of chronic lead
    poisoning in Jiansu Province, China, involving people who had drunk
    rain water boiled in tin kettles. Analysis of the lead content of
    water showed that after boiling the water contained 0.79 to
    5.34 mg lead/litre.

    5.3  Miscellaneous exposure

    5.3.1  Cosmetics and medicines

         Some traditional medicines and customs have been found to result
    in exposure to high levels of lead, most of which cannot be quantified
    with any degree of accuracy. Rather than occurring as trace
    ingredients or trace contaminants, various lead compounds are used as
    major ingredients in traditional medicines in numerous parts of the
    world (Table 18). Clinically overt lead poisoning due to traditional
    cosmetics and medicines has been identified among infants (Shaltout et
    al., 1981; Fernando et al., 1981; Sharma et al., 1990), children and
    adults (Pontifax & Garg, 1985; Cueto et al., 1989; Mitchell-Heggs et
    al., 1990; Gupta et al., 1990). There are case reports of lead
    toxicity secondary to inhalation of lead from traditional remedies
    (Aslam et al., 1979; Shaltout, 1981; Cueto et al., 1989; Sharma et
    al., 1990; Mitchell-Heggs et al., 1990).

         Often the use is not limited to adults; these may be used on
    infants and young children, as well as on women. In Kuwait, the leaded
    "kohl", also called "Al kohl", is traditionally applied to the raw
    umbilical stump of the newborn in the erroneous belief of a beneficial
    astringent action (Fernando et al., 1981). An additional use of lead
    metal and lead sulfide is for inhalation of the fumes ("Bokhoor")
    produced from heating on hot coals, in the mistaken belief that this
    will calm irritable infants and children (Fernando et al., 1981;
    Shaltout et al., 1981).

         Latin-American countries also report the use of traditional
    medicines with high lead concentrations. For example, the Mexican
    traditional remedy "azarcon" (lead chromate) and/or "greta" (mixed

    lead oxides), distributed as finely ground powders, may contain more
    than 70% lead. They are used in the treatment of "empacho", a
    gastrointestinal disorder considered to be due to a blockage of the
    intestine (Trotter, 1990).

         In addition to the potential risks of lead exposure from the use
    of traditional medicines, clinical lead poisoning can result from the
    lodging of lead shot  in vivo (Manton & Thal, 1986).


        Table 18.  Sources of lead exposure in traditional medicines and cosmetics
                                                                                      
    Source of lead   Comments                                   Reference
    (product)
                                                                                      

    Summa/Kohl       used in Indo-Pakistan and other Muslim     Aslam et al. (1979);
                     cultures as eyes preparation; placed on    Fernando et al. (1981);
                     conjunctival surface or as astringent on   Shaltout et al. (1981);
                     umbilical cord stump. Antimony originally  Sharma et al. (1990)
                     used but lead cheaper.

    Hindu folk       ground seeds and roots as treatment        Pontifax & Garg (1985)
     medicine        for diabetes (8 mg lead/g)

    Bokhoor          tribal custom to produce lead fumes to     Shaltout et al. (1981)
                     ward off evil

    Azarcon          lead chromate and mixed lead oxides as     Trotter (1990)
                     treatment for gastrointestinal disorders
                     in Mexico and southwestern USA

    Skin ointments   cosmetics used by Chinese actors; skin     Lai (1977)
     and cosmetics   ointment in Europe
                                                                                      
    

    5.4  General population exposure

         The total intake of lead by adults and children in the general
    population varies greatly as to the relative contributions from
    individual sources (air, water, food, soil/dust and others) and is
    partly dependent on life-style and socioeconomic status. It is beyond
    the scope of this review to provide comprehensive information covering
    a wide range of circumstances. However, a few simplified calculations
    will be given as guidance for carrying out such determinations.

         Table 19 gives a summary of the total lead intake and uptake from
    the general environment in adults and in children aged 1 to 5. The
    assumptions made are shown and are taken from WHO (1987). Additional
    intake of lead will take place in certain groups from the use of
    tobacco and alcoholic beverages. The estimates given are probably on
    the high side with respect to the contribution from air and dust,
    since indoor and outdoor lead contributions were considered equal.

         Additional intake of lead is possible due to ciliary clearance of
    particles 1-5 µm in diameter with subsequent swallowing and
    gastrointestinal absorption.

    5.5  Blood lead concentrations of various populations

         Under certain conditions, blood lead (PbB) levels are a useful
    indicator of exposure and are therefore discussed here, as well as in
    section 6.1.4. In general, PbB levels correlate better with recent
    exposure levels (Lyngbye et al., 1990b). As is discussed in sections
    5.5.1 and 5.5.2, the general trend observed in all blood lead surveys
    carried out in countries engaged in risk reduction programmes over the
    last 15 to 20 years is a fall in the measured levels (OECD, 1993).

         PbB concentrations are the most often used estimate of general
    exposure to lead. The USA, Commission of European Communities (CEC),
    Australia and the WHO have carried out epidemiological surveys to
    determine lead exposures in various populations. The designs of these
    studies differed; some provided estimates for the country as a whole,
    others compared PbB levels of people working in the same occupation in
    various countries, and others surveyed general populations but were
    not designed to be extrapolated as national estimates. What these
    studies have in common is that they did not emphasize groups with high
    lead exposures, but concentrated on typical PbB concentrations of
    non-occupationally-exposed groups in the regions or countries studied.
    These surveys also offer the possibility of examining the data as
    distributions, especially at the high end of the distribution profile.
    Such examination may identify unusual exposures and thus populations
    at risk. Average PbB concentrations may disguise the risk to various
    segments of the population.

         In the USA, national estimates of the extent and severity of
    recent human exposures to lead in the general population were based on
    PbB measurements from the second National Health and Nutrition
    Examination Survey of 1976-1980 (NHANES II) (Annest & Mahaffey, 1984).
    The design of this study permitted extrapolation to the USA population
    as a whole.

        Table 19.  Estimates of lead (µg/day) absorbed by adults and children
               from air, dust, food and watera
                                                                                 
    Mean air       Dust                 Source of lead (µg/day)           Total
    lead           intake                                               absorbed
    concentration  (mg/day)b                                            (µg/day)
    (µg/m3)                       Air       Dust      Food      Water
                                                                                 

    Adults

        0.3           N.S.        2.4        --        10          2        14.4
        0.5           N.S.        4.0        --        10          2        16.0
        1.0           N.S.        8.0        --        10          2        20
        2.0           N.S.       16.0        --        10          2        28

    Children 1-5 years

        0.3             --        0.6        --        25          5        30.6
        0.5             --        1.0        --        25          5        31.0
        1.0             --        2.0        --        25          5        32.0
        2.0             --        4.0        --        25          5        34.0
        1.0             25        2          12.5      25          5        44.3
        1.0             50        2          25.0      25          5        57.0
        1.0            100        2          50.0      25          5        82.0
        1.0            200        2         100.0      25          5       132.0
                                                                                 

    a  Adapted from WHO (1987); Dust is not considered a significant
       source of lead in adults, but is a significant source for workers
       where hygiene practices are poor

       The above estimates are based on the following assumptions:

       Air: Respiratory volume in adults is 20 m3/day, and in children
       5 m3/day, and the respiratory absorption is 40%.

       Food: Intake of lead by adults 100 µg/day with 10% absorption and
       50 µg/day for children with 50% absorption.

       Water: A lead concentration of 20 µg/litre, with adult consumption
       of 1 litre/day and 10% absorption and for children 0.5 litre/day
       with 50% absorption.

       Dust: Dust concentration of lead was 1000 µg/g and absorption was
       50%.

    b  N.S. = Not significant
    
         The CEC study was directed toward biological screening of the
    extent of exposure to lead outside the workplace using PbB level as
    the index of exposure (CEC, 1981). The object of the study was to
    assess non-occupational exposure of the population to lead in the
    member states; it was implemented by member states and coordinated by
    the CEC. Specific populations studied were selected by member states
    and 168 separate areas and population groups were investigated in
    1977.

         Two other major international studies were limited to urban
    populations. A pilot study of human lead exposure (Friberg & Vahter,
    1983), organized by the United Nations Environment Programme and WHO,
    was a collaborative effort under the Global Environmental Monitoring
    System (GEMS). The following countries participated in at least part
    of the study: Belgium, China, India, Islamic Republic of Iran, Israel,
    Japan, Mexico, Peru, USA and the former Yugoslavia. The subjects of
    the study were teachers, since they comprise an occupational group not
    extensively exposed to lead. An Australian study was based on urban
    residents only, and included 651 subjects between 6 and 91 years of
    age (Hopper et al., 1982).

         Based on demographic, economic and individual variables found in
    the USA study to be associated with PbB levels (Mahaffey et al.,
    1982), it is clear that results from these major studies cannot be
    directly compared. However, a comparison of such studies can provide
    information on the exposure to lead of segments of the various
    populations in various countries worldwide.

    5.5.1  Adult populations

         In a large health screening programme within the USA during
    1976-1980 (NHANES II), over 27 000 residents aged 6 months to 74 years
    were examined (Annest & Mahaffey, 1984). PbB concentrations were
    determined in a subsample (9933 individuals) and yielded an arithmetic
    mean value of 0.67 µmol/litre (13.9 µg/dl). In 5841 individuals aged
    18-74 years, the overall mean was 0.68 µmol/litre (14.1 µg/dl)
    (Roberts et al., 1985). A sex difference was noted: 0.77 µmol/litre
    (16.1 µg/dl) was found in males and 0.57 µmol/litre (11.9 µg/dl) in
    females. Residents in the centre of large urban areas were found to
    have a PbB level of 0.72 µmol/litre (14.9 µg/dl), while rural
    residents it was 0.62 µmol/litre (13.0 µg/dl). Because of the design
    of the study, information on PbB levels, in relation to smoking,
    alcohol consumption, and occupational status, were available. Male
    workers in occupations with a high potential for lead exposure had
    mean PbB levels of 0.78 µmol/litre (16.2 µg/dl) for nondrinkers/
    non-smokers, 0.92 µmol/litre (19.2 µg/dl) for non-drinkers/non-smokers
    and 0.95 µmol/litre (19.7 µg/dl) for drinkers/smokers. It was noted
    that a true decrease of 37% in PbB level occurred during the period of
    the survey (Annest et al., 1983).

         Rabinowitz & Needleman (1982) reported an arithmetic mean
    umbilical cord PbB concentration of 0.32 µmol/litre (6.6 µg/dl) (range
    0-1.78 µmol/litre; 0-37 µg/dl) in over 11 000 samples collected
    between 1979 and 1981. A decrease in PbB level of approximately 11%
    was noted during the period of collection.

         Friberg & Vahter (1983) reported the PbB levels of 200 teachers
    from nine countries (see section 5.5). The median PbB level in this
    study ranged from 0.29 µmol/litre (6 µg/dl) in Beijing and Tokyo to
    1.06 µmol/litre (22 µg/dl) in Mexico City. Smokers generally had
    higher PbB levels than non-smokers.

         During the period 1978-1988 marked decreases (approximately 30 to
    40%) in the average PbB levels of adults were noted in Belgium,
    Germany, New Zealand, Sweden, the United Kingdom, and the USA (OECD,
    1993).

    5.5.2  Children

         Results from the NHANES II survey in the USA indicated a mean PbB
    level for children (6 months to 2 years old) of 0.78 µmol/litre
    (16.3 µg/dl). It was noted that 18.6% of black inner city children had
    PbB levels > 1.44 µmol/litre (30 µg/dl) whereas only 4.5% of white
    children had such PbB concentrations. The percentage of children with
    high PbB level decreased with increasing family income (Roberts et
    al., 1985). The NHANES II data were analysed in detail for information
    on the demographic correlates of children's PbB concentrations
    (Mahaffey et al., 1982; Annest et al., 1983). The racial differences
    in these proportions were significant among both boys and girls. The
    proportion with elevated PbB concentrations was slightly higher among
    boys than girls for both races. Among both black and white children,
    the percentage with elevated PbB concentrations decreased with
    increasing family income. Proportionately more young children in urban
    than rural areas and in the central cities of large urban areas had
    elevated blood lead concentrations.

         PbB concentrations were measured in 286 Finnish children living
    in the three largest cities of Finland (number of subjects = 172), in
    rural areas (54 subjects) and in a lead smelter area (60 subjects)
    (Taskinen et al., 1981). The mean PbB levels in urban, rural and
    lead-smelter areas varied between 0.29 µmol/litre and 0.32 µmol/litre,
    range 0.14-0.82 µmol/litre (6.0 and 6.7 µg/dl, range 3-17 µg/dl).
    There were no significant differences between the areas of residence.
    The five children who lived within 500 metres of the lead smelter
    had a mean PbB concentration of 0.44 µmol/litre, range 0.24-0.62
    µmol/litre (9.2 µg/dl, range 5-13 µg/dl), which was significantly
    higher than the mean blood lead concentration among children living
    in the rest of the country.

         In a study carried out in Sweden, which included 1781 samples
    obtained from children during 1978-1988, the mean PbB level decreased
    from 0.29 µmol/litre (59.6 µg/litre), range 0.09-1.20 µmol/litre
    (18-250 µg/litre) in 1978 to 0.16 µmol/litre (32.9 µg/litre), range
    0.07-0.34 µmol/litre (15-71 µg/litre) in 1988 (Schutz et al., 1989).
    In Finland, a remarkably similar decrease has been reported; the mean
    PbB value was 0.14 µmol/litre (30 µg/litre), range 0.1-0.40 µmol/litre
    (21-41 µg/litre) among 35 children in the Helsinki area in 1988 (Ponka
    et al., 1991).

         As in the USA (Annest et al., 1983), children's PbB
    concentrations in Sweden have decreased in recent years. Levels were
    determined each summer during the period 1978-1984 in children from
    Scania in Southern Sweden (Skerfving et al., 1986). The average PbB
    concentration was 0.26 µmol/litre, range 0.07-1.20 µmol/litre
    (5.5 µg/dl, range 1.4-25.0 µg/dl). There was a statistically
    significant decrease over time in both the rural and urban areas,
    averaging about 0.019 µmol/litre (0.4 µg/dl) per year.

         Decreases of 25-45% in average PbB levels in children between
    1978 and 1988 have been reported in Belgium, Canada, Germany, New
    Zealand, Sweden and the United Kingdom (OECD, 1993).

    5.5.3  Remote populations

         In contrast, a much higher mean value of 0.16 µmol/litre
    (3.4 µg/dl) was obtained in Nepal upon examination of 103 children
    living in the Manang District, which is also very remote and thought
    to be free of anthropogenic sources of lead (Piomelli et al., 1982).
    However, these people are exposed to significant amounts of natural
    lead from living in dusty smoke-filled houses, and from burning pine
    and yak dung. The lead level in the houses was about 0.15 µg/m3
    (Davidson et al., 1981). For comparison outdoor air in Woods Hole,
    Massachusetts, USA, contains 0.004 and Boston, Massachusetts, 0.01 to
    0.05 µg lead/m3.

         Another remote population studied was from Okapa, Eastern
    Highlands, Papua New Guinea (Poole et al., 1980). Extreme care was
    taken in sampling and analysis, and the precision was acceptable at
    PbB levels as low as 0.24 µmol/litre (5 µg/dl). The children subsisted
    on root crops and a small amount of tinned fish and meat. Villages
    were typically hazy with wood smoke, natural mineral pigments were
    used extensively, and tobacco smoking was common. This survey of 100
    children yielded a mean PbB level of 0.25 µmol/litre (5.2 µg/dl), well
    above the detection limit, and, as reported by Poole et al. (1980),
    far less than the mean of 1.10 µmol/litre (23 µg/dl) for Sydney,
    Australia, in 1974.

    5.6  Occupational exposure

         Occupational exposure to lead which results in poisoning, both
    moderate and clinically symptomatic, still occurs in many countries of
    the world. Although adults are mainly involved, in many countries,
    especially in those with developing industries and small home-based
    industries, the distinction between home and workplace is non-existent
    (Verrula & Noah, 1990) and children are exposed to workplace lead.
    Until recently, in the occupational setting, there was only concern
    for the identification of late stage, highly symptomatic cases of lead
    poisoning resulting in crippling neurological manifestations of lead
    poisoning such as palsy and encephalopathy (Chakravorti & Bhar, 1978;
    Wang, 1984; Davies, 1984; Lin-Fu, 1985). However, there is now concern
    for lower exposures to lead.

         In many countries occupational lead exposure is entirely
    unregulated and no monitoring of exposures exists. Automobile battery
    manufacture and repair, radiator repair, secondary smelters (including
    scrap metal refiners) are found in most countries. The industries
    where there is a potential for lead exposure are listed in Table 10
    (section 5.1.3). Significant lead exposures are not limited to
    traditional heavy industries. For example, Kaye et al. (1987)
    identified exposures from leadborosilicate dust used in a capacitor
    and resistor plant in the USA.

         Because of transfer of lead to the fetus  (in utero) and the
    transport to the home of lead on clothing, etc., thereby exposing the
    young child in the home, the problems of occupational exposures to
    lead are not limited to the workplace  per se. Cases of poisoning
    among the children of lead workers have been reported (Baker et al.,
    1979; Wang, 1984; Ryu et al., 1985).

         Although most occupational standards are based on airborne lead
    only, this route of exposure does not fully reflect the total daily
    exposure of workers, also exposed to lead in food, water, alcoholic
    beverages and dusts.

         The potential for hazardous exposures to lead during lead
    smelting and refining are well recognized, particularly where molten
    lead and alloys are poured, resulting in the vaporization of metal.
    This is true for both primary new metal and secondary (lead scrap)
    smelters and refineries.

         Small domestic versions of secondary smelters exist in a large
    number of countries. These are typically located within or in close
    proximity to homes. For example, in Jamaica there has been a rapid
    proliferation of lead smelters, particularly illegal backyard smelters
    (Rodney & Lee, 1985). The lead fumes and dust generated in such
    operations pose an exceptional health hazard to children and adults

    living near these operations. Rodney & Lee (1985) reported that 51% of
    116 children (aged 2 to 12 years, mean 5.9 years) and 60% of 235
    adults working in or living near lead smelting factories had PbB
    concentrations of 1.92 µmol/litre (40 µg/dl) or more.

         Other occupations where workers have been shown to be at risk
    from airborne lead include electric storage battery manufacturing,
    particularly where industrial hygiene is poor (Barrio & Badia, 1985);
    demolition, welding and shipbreaking where lead-based paint is present
    (Holness & Nethercott, 1988; Nosal & Wilhelm, 1990); pottery and
    ceramic-ware production, which is often a home-based operation
    involving women and children (MolinaBallesteros et al., 1983; Katagiri
    et al., 1983; Kaye et al., 1987); small businesses repairing
    automobile radiators (Matte et al., 1989a,b; Verrula & Noah, 1990),
    and artisans producing jewellery and decorative wares. This latter
    industry is of particular concern since it is predominantly carried
    out at home or in non-regulated shops, often by women and children
    (Behari et al., 1982). Indian silver jewellery makers were found to
    have a PbB level of 5.8 µmol/litre (121 µg/dl) compared to non-exposed
    controls with a PbB level of 1.3 µmol/litre (27 µg/dl) (Behari et al.,
    1982).

    6.  KINETICS AND METABOLISM IN LABORATORY ANIMALS AND HUMANS

    6.1  Absorption

         The absorption of lead from environmental sources is not
    dependent solely on the amount of lead presented to the portals of
    entry. It is also dependent on the physical and chemical state in
    which the metal is presented, and it is influenced by host factors
    such as age, physiological status, nutritional condition and,
    possibly, genetic factors. Men engaged in heavy work breathe more air
    and eat more food than sedentary individuals of the same weight, and
    on a body weight basis children eat almost as much food and breathe
    almost as much air as middle-aged adults.

         Considerable data from human subjects are available. Therefore,
    discussion of animal studies will be limited to areas where the
    information on humans is inadequate. Various models used to predict
    body burden or distribution of lead have been developed (Bernard,
    1977; Marcus, 1985a,b,c; Bert et al., 1989; Arnetz & Nicolich, 1990).
    A full review of these models has been judged to be beyond the scope
    of this document. However, some of them were used in section 6.1.4
    when correlating intake and body burdens.

    6.1.1  Absorption after inhalation

         The absorption of lead from air to blood involves two processes:
    the deposition of airborne lead particles in the respiratory tract;
    and the absorption and clearance from the respiratory tract into the
    circulation. Using the International Radiological Protection
    Commission (IRPC) document on lung dynamics (Task Group on Lung
    Dynamics, 1966), a model was developed which predicted that 35 to 50%
    of inhaled lead is deposited in the respiratory tract (40-50% of
    particles with a mean mass median aerodynamic diameter (MMAD) of
    0.5 µm, such as are typically generated by automobiles). These are
    deposited primarily in the alveolar sacs of the lung. Lead fumes and
    vapours, such as those generated in operations where metals are cut or
    heated, are of very small size and are respirable. Absorption after
    deposition will vary according to the solubility of the lead species
    (e.g., lead carbonate or lead chloride aerosols) and the inherent
    toxicity to lung macrophages and cilia.

    6.1.1.1  Animal studies

         Limited animal studies confirm that there is almost complete
    absorption of lead particles (0.1 to 0.5 µm in diameter) deposited in
    the lower respiratory tract. In rats the clearance half-time from lung
    is short (less than one hour) and 90 to 98% of the administered dose
    is absorbed within about 48 h (US EPA, 1986a).

    6.1.1.2  Human studies

         The respiratory deposition of airborne lead is in the range of
    30-50% and varies with particle size and ventilation rate (US EPA,
    1986a). Higher deposition rates may occur with larger particles, but
    this deposition takes place in the upper respiratory tract, with
    eventual displacement to the gastrointestinal tract and absorption via
    the ingestion route. This probably explains the observation of Kehoe
    (1961) that faecal excretion of lead increased after a subject
    breathed for many weeks aerosols of lead oxide (150 µg/m3) with an
    MMAD of approximately 2.9 µm. In contrast, smaller particles of
    inhaled lead, such as those generated by automobile exhaust,
    regardless of physicochemical form, are almost (> 90%) completely
    absorbed after deposition in the lower respiratory tract (Rabinowitz,
    et al., 1977a; Chamberlain et al., 1978; US EPA, 1986a).

         There are no quantitative data on the absorption of lead in
    children after inhalation exposure. It is known that young children
    weighing only one sixth of an adult inhale 40% of the daily volume of
    an adult and a proportionately higher daily air volume per unit
    measure (weight, body area) than do adults (Barltrop, 1972). After
    controlling for weight and taking into account differences in the
    anatomy of the respiratory tract between adults and children, James
    (1978) calculated a rate of deposition of lead particles in children
    which was 1.6 to 2.7 times that of adults.

    6.1.2  Absorption of lead from the gastrointestinal tract

         In the case of older children and adults without occupational
    exposure, lead absorbed by the gastrointestinal tract comes from the
    intake of lead in foods, beverages and soil/dust. In pre-school
    children, there is concern over the intake of both food and non-food
    items (e.g., toys, soil/dust). Young children may take in lead from
    non-food items, via normal mouthing activity, which in the extreme, is
    the behavioural trait pica, which refers to the ingestion of such
    materials as soil, ash, paint chips and plaster (US EPA, 1986a). For
    infants and young children, the extent of absorption of the lead in
    dust/soils from the gastrointestinal tract is extremely important,
    particularly for children living in urban environments.

    6.1.2.1  Animal studies

         The absorption of lead from the gastrointestinal tract in
    experimental animals is age dependent and is modified by the level of
    food intake. In 1-week-old suckling rats, an absorption rate of about
    52% was reported after a single oral dose of lead chloride, compared
    to 0.4% in 6-week-old animals (Kostial et al., 1978). Furthermore,
    Aungst et al. (1981) reported that rat pups had higher tissue levels
    of lead than adults after a single gavage dose of 1 or 10 mg lead (as

    lead acetate) per kg body weight. Fasting markedly enhanced the uptake
    of lead in the gastrointestinal tract of experimental animals (Garber
    & Wei, 1974; Pounds et al., 1978).

         In experimental animals, absorption of lead from the
    gastrointestinal tract appears to be a saturable process. With
    increasing doses (1 to 100 mg lead (as lead acetate) per kg body
    weight), lead absorption as a percentage of dose was found to decrease
    from 42% to 2% (Aungst et al., 1981; Bushnell & DeLuca, 1983) in
    dietary studies. Such data are consistent with a saturable active
    transport process across the gastrointestinal tract.

         The low rate of absorption of lead from soluble lead salts noted
    in adult rats may reflect a dietary effect. Kostial & Kello (1979)
    reported less than 1% absorption in adult rats given chow diets,
    whereas the absorption rate was 3-20% when rats consumed diets similar
    to that of humans (milk, bread, baby foods, etc). This range more
    closely resembles the human absorption rates.

         The chemical form of lead can influence its bioavailability.
    Stone et al. (1981) determined the biological availability of lead
    intrinsically incorporated into the soft tissues of oysters. The
    bioavailability to the rat was 10-30% lower than that of lead acetate
    added to the basal chow diet. Henning & Cooper (1988) compared the
    absorption of lead from rat milk labelled  in vitro with lead-203 and
    lead chloride or lead acetate solutions. Lead from soluble salts
    accumulated primarily in the duodenum, to some extent in the jejunum
    and minimally in the more distal small intestine. Lead from milk
    accumulated only in the upper ileum. After 20 h negligible lead-203
    was found in any region following administration of soluble salts, but
    there was substantial retention of lead in the ileum after dosing with
    milk.

         Data obtained from a feeding study in rats (Dieter et al., 1993)
    showed that lead uptake into rat femurs was highly dependent on the
    chemical form of lead administered. Bioavailability was highest for
    lead acetate, intermediate for lead oxide, and lowest for lead sulfide
    and Alaskan mixed ore concentrate. This uptake was linearly related to
    dose over the range studied. However, the slopes of the linear
    regression equations differed according to the form of lead. They were
    0.10 µg lead/g femur per kg diet for lead sulfide and 2.64 for lead
    acetate.

         However, results of an investigation in young swine of the
    absorption of lead from mining and/or milling operation waste (LaVelle
    et al., 1991) differ from those reported by Dieter et al. (1993). The
    chemical species (predominantly lead sulfide) and particle size (i.e.
    larger than 100 µm) were expected to result in these sources being
    less bioavailable than soluble lead salts such as lead nitrate.
    However, experimental data indicate that the lead in mine tailings may
    be about 2-3 times more available than reagent grade lead sulfide

    under the conditions of the study. Additionally, Freeman et al. (1992)
    reported that the lead in mining waste soil was between 8 and 20% as
    bioavailable as lead acetate. No detailed comparisons were made of the
    physicochemical properties of these two mine wastes. An additional
    study involving the absorption of lead from lead sulfide ores and
    their oxidation products in rabbits and  in vitro showed that lead
    from minerals was absorbed less well than from lead acetate by a
    factor of 5 (Davis et al., 1990a). The significance of these phenomena
    is not clear.

    6.1.2.2  Human studies

         Gastrointestinal absorption of lead in humans, as in experimental
    animals, is influenced by dietary factors, nutritional status and the
    chemical form of the metal. Overall patterns of food intake may also
    influence lead absorption. For example, lead ingested during periods
    of fasting is absorbed to a much greater extent than lead ingested
    with food. Chamberlain et al. (1978) reported 45% absorption of lead
    chloride in fasting subjects and only 6% in feeding subjects. Using a
    similar procedure, Heard & Chamberlain (1982) reported absorption of
    63.3% in fasting subjects. Using multiple stable isotope lead tracers,
    it was found that in adult men the gastrointestinal absorption rate
    with food containing lead nitrate or lead cysteine was 6-12%. However,
    when consumed under fasting conditions, lead nitrate, lead sulfide and
    lead cysteine were absorbed at 16-53% (Rabinowitz et al., 1977b).

         Kehoe (1961) estimated only 10% net absorption of dietary lead by
    adults in long-term metabolic studies. However, with appropriate
    calculation of biliary clearance as well as of urinary excretion, a
    figure of 15% was estimated by Chamberlain et al. (1978). In several
    studies on adult humans, absorption of lead was reported to be 14%
    when it was administered with food (Chamberlain et al., 1978; Moore et
    al., 1979; Rabinowitz et al., 1980).

         Ziegler et al. (1978) reported that young children, aged two
    weeks to two years, absorbed 42% of ingested lead at levels of intake
    greater than 5 µg/kg body weight. Drill et al. (1979) estimated an
    absorption rate of 17% for lead in paint chips in children aged 2 to 3
    years. These authors also estimated a 30% gastrointestinal absorption
    rate for lead in soil and dirt.

         The amount of lead ingested by children from non-food items such
    as soil, dust and paint chips through normal mouthing activity,
    particularly for children with pica, is a major concern in calculating
    paediatric lead exposures. Using data from Day et al. (1975) and Lepow
    et al. (1974), it has been estimated that children 2 to 3 years of age
    ingest about 100 mg soil per day. Using aluminium and titanium as
    tracers, Clausing et al. (1987) estimated that children aged 2 to 4
    years ingest between 50 and 100 mg of soil daily. The average amount
    of soil ingested by young children has recently been estimated to be
    between 12.5 and 21 mg/day (SAHC, 1993).

    6.1.2.3  Nutritional status and lead absorption via gastrointestinal
             tract

         It has been known for some time that the absorption and
    distribution of lead are affected by nutritional status in both
    experimental animals and humans (Sobel et al., 1940). Nutritional
    inadequacies can also affect the toxic response to lead (see sections
    7 and 8). In view of documented nutritional inadequacies in many parts
    of the world, such interrelationships become crucial in assessing the
    full risk from lead exposures.

         Vitamin D, calcium and phosphorus have complex and interrelated
    effects on lead absorption (Fullmer, 1990). Increasing the
    concentration of 1,25-dihydroxycholecalciferol, the active metabolite
    of vitamin D, either exogenously or endogenously, increases
    gastrointestinal absorption of lead (Fullmer, 1990). However, this
    effect is dependent upon the duration of lead exposure and the
    magnitude of body lead stores. This homoeostatic mechanism for calcium
    and its dependence on nutritional status, as well as body burden of
    lead, is complex. This may explain the divergent results of the
    observed interaction in children (Rosen et al., 1980) or the lack of
    association (Koo et al., 1991) of lead with vitamin D metabolism.

         In experimental animals, chronic ingestion of diets with less
    than adequate amounts of calcium (Quarterman & Morrison, 1975),
    phosphorus (Quarterman & Morrison, 1975), iron (Mahaffey-Six & Goyer,
    1972; Ragan, 1977), selenium (Stone & Soares, 1976) or zinc
    (Cerklewski & Forbes, 1976) increases the fractional absorption of
    lead in the gastrointestinal tract. From the work of Fullmer & Rosen
    (1990), it appears that lead alters the transport of calcium and other
    trace minerals by affecting carrier proteins, rather than by competing
    at the mucosal surface.

         The influence of iron on lead absorption in humans was assessed
    in double-labelling experiments conducted by Watson et al. (1980).
    Lead-203 and iron-59 were given to 28 subjects in their diet. There
    were significant positive correlations between iron levels in food and
    lead absorption. However, the association was not strong because 50%
    of those absorbing excess lead also absorbed excess iron.

         Rats fed iron-deficient diets have increased concentrations of
    lead in kidney and bone when compared to rats ingesting equivalent
    quantities of lead (as lead acetate) in drinking-water while being
    fed an iron-adequate diet (Mahaffey-Six & Goyer, 1972). In contrast
    with calcium deficiency, iron deficiency does not result in a
    redistribution of lead to non-osseous tissue (Mahaffey-Six & Goyer,
    1972). The degree of iron deficiency does not need to be severe to
    increase retention of lead. For example, Ragan (1977) demonstrated
    six-fold increases in tissue lead in rats when body iron stores were
    reduced, but before frank iron deficiency developed. High levels of
    dietary iron resulted in decreased kidney, femur and blood lead

    concentrations in rats (Mahaffey, 1985). Iron appears to increase
    absorption rather than decrease lead excretion (Barton et al., 1978).
    Lead interferes with normal transferrin transport of iron, and
    inhibits transferrin endocytosis and iron transport across the cell
    membrane of the reticulocyte (Qian & Morgan, 1990).

    6.1.3  Dermal absorption

    6.1.3.1  Human dermal absorption

         Moore et al. (1980) examined the uptake of lead acetate from two
    hair-darkening cosmetics through the skin of eight human volunteers.
    Only minute quantities of lead (0-0.3% of the applied dose) were
    detectable in blood, and there was only a slight increase in
    absorption when the skin was damaged. Lilley et al. (1988) and
    Florence et al. (1988) have reported the dermal absorption of
    inorganic lead compound leading to elevated levels of lead in human
    saliva and sweat.

         Dermal absorption of inorganic lead through unabraded human skin
    is considered to be minimal.

    6.1.4  The relationship of external lead exposure to blood lead
           concentration

         There are numerous environmental media that provide routes by
    which humans are exposed to lead: food, water, air, soil, dust.
    External exposures are the sum of the quantities of lead consumed from
    all sources.

         Historically there have been two lines of approach in
    understanding the lead exposure and blood lead relationship. Most have
    been empirical and have measured environmental lead and PbB levels
    either at one time or repeatedly. With these observed correlations and
    with no assumptions about how lead moves inside the body, many
    reasonable predictions can be made. The validity of these predictions
    is optimum when only one environmental source dominates. Some of these
    rely on linear functions, while others have specified non-linearities,
    especially over very wide ranges (more than a factor of 10) of lead
    exposures. However, when multiple sources are considered these
    predictive models have been less satisfactory.

         Internal lead levels in human populations have been estimated by
    analyses of a variety of biological tissues (e.g., blood, teeth, bone
    and hair). Lead concentrations in each of these have particular
    biological meanings with regard to external exposure to lead. Blood is
    the compartment in which lead is most often measured as a marker of
    exposure. However, it typically represents relatively recent
    exposures, since the half-life of lead in blood is short (US EPA,
    1986a) and has been estimated to be in the order of 36 days from
    tracer studies (Rabinowitz et al., 1975). Lead in blood is derived

    from levels in the current environment and from lead stored in tissues
    (particularly bone) that re-enters the blood during tissue
    mobilization (Manton, 1985). Although PbB concentration reflects
    recent exposure and bears a consistent relationship to levels of lead
    in the external environment, when bone mobilization is accelerated a
    greater fraction of PbB will be derived from tissue stores.

         To date only a few studies have utilized a multimedia approach
    relating lead intake to PbB levels. The majority of studies have
    attempted to correlate PbB levels and lead concentrations in specific
    media.

    6.1.4.1  Ambient air

    a)   Occupational exposure

         Several studies have examined the blood lead: air lead
    relationship for workers exposed to levels of airborne lead between 9
    and 450 µg/m3 (King & Conchie, 1979; Gartside & Buncher, 1982;
    Bishop & Hill, 1983). The results of all three studies are in general
    agreement over the wide range of airborne lead studied and for PbB
    levels between 0.96 and 4.32 µmol/litre (20-90 µg/dl). The blood lead:
    air lead relationship in occupational settings is best described by a
    curvilinear relationship having slopes between 0.00096 and
    0.0038 µmol/litre (0.02 and 0.08 µg/dl) per µg/m3 air.

    b)   Non-occupational exposure

         Both population and experimental studies have been used to
    estimate the PbB: ambient air lead relationship in adults and
    children. Under ambient conditions (air lead concentrations of
    0.1-2.0 µg/m3) and PbB levels less than 1.44 µmol/litre
    (< 30 µg/dl), the relationship has been described as linear (Colombo,
    1985). The various slope estimates reported are based on the
    assumption that an equilibrium level of lead in blood is achieved.
    Based on three major studies (Yankel et al., 1977; Angle & McIntire,
    1979; Roels et al., 1980), the median slope for children is about
    0.091 µmol/litre (1.9 µg/dl) blood per µg/m3. In adult males a slope
    estimate of 0.076 µmol/litre (1.6 µg/dl) blood per µg/m3 was
    calculated. When one calculates the relationship between PbB and the
    total contribution from air (direct inhalation plus indirect through
    dust/soil), a value of about 0.14-0.24 µmol/litre (3-5 µg/dl) blood
    per µg/m3 is obtained (Brunekreef, 1984; US EPA, 1986a). Also on the
    basis of a linear model, Snee (1981) reported that the best estimate
    of the blood lead: air lead relationship was 0.048 µmol/litre
    (1 µg/dl) per µg/m3. After examining available epidemiological and
    experimental data, Chamberlain (1983) concluded that most published
    estimates of the slope were between 0.072 and 0.144 µmol/litre per
    µg/m3. From these results, it can be concluded that airborne lead
    will only be a major contributor to PbB levels in areas of high air
    lead levels.

    6.1.4.2  Food

         The relationship of PbB to dietary intake has been estimated from
    experimental (Stuik, 1974; Cools et al., 1976; Schlegel & Kufner,
    1979) as well as population studies (Sherlock et al., 1982; UK, 1982;
    Ryu et al., 1983). In adults, the results from both types of studies
    using both linear and cube root models indicated a relationship of
    between 0.0019 and 0.0028 µmol lead/litre (0.04 and 0.06 µg/dl) per µg
    lead intake per day. From the study of Ryu et al. (1983) a slope of
    0.0096 µmol/litre (0.2 µg/dl) per µg lead/day was obtained for infants
    aged 8 to 196 days.

         Currently, data are available for adults and children from
    studies with careful control of important variables such as: intake of
    dietary lead and of other dietary constituents, minimal exposure to
    sources other than diet in the studies of infants, and intake/blood
    lead measurements that can be used to estimate intake from all
    sources. For infants these are studies reported by Sherlock et al.
    (1982), UK (1982), Ryu et al. (1983), and Lacey et al. (1985). Studies
    by the UK (1982) and by Sherlock et al. (1982) were conducted at
    higher levels of lead exposure than were the studies conducted with
    infants by Ryu et al. (1983). The latter studies were at exposure
    levels associated with PbB concentrations under 0.96 µmol/litre
    (20 µg/dl); the ratio of blood lead to ingested lead was
    0.0076 µmol/litre (0.16 µg/dl) per µg lead ingested per day. Exposure
    levels were low: average dietary lead intake was 17 µg/day. After four
    months, the average PbB concentration was 0.29 µmol/litre (6.1 µg/dl)
    whole blood. Between the ages of 4 and 6 months, 10 children remained
    at a dietary lead intake of 0.76 µmol/litre (16 µg/dl). Their PbB
    concentrations were quite constant at 0.35 µmol/litre (7.2 µg/dl) at
    the end of the study period. The remaining seven children received
    canned infant formula and/or milk during this period and their average
    dietary lead intake was 61 µg/day. At the end of the study period,
    their PbB concentration was 0.69 µmol/litre (14.4 µg/dl). Based on
    these data, a curvilinear relationship between blood lead and total
    lead intake was suggested, a 4-fold increase in lead intake resulting
    in a doubling of PbB concentration. Sherlock et al. (1982) conducted a
    duplicate diet study of 31 mothers and their children from Ayr,
    Scotland. The slope for adults was substantially lower than for
    children.

    6.1.4.3  Drinking-water

         There is still debate over the most appropriate model (i.e. cube
    root, polynomial or logarithmic) to describe the curvilinear
    relationship of waterborne lead to blood lead. The highest estimates
    for the contribution of water lead to blood lead come from the cube
    root and logarithmic models. Much lower estimates are obtained from a
    linear model (Pocock et al., 1983; US EPA, 1986a). In a study by
    Sherlock et al. (1982), a cube root relationship between lead levels

    in drinking-water and blood fitted the data more closely than a linear
    relationship. The curvilinear model implies that as abatement
    processes lower water lead concentrations, there will be an increasing
    benefit in lowering of population PbB levels (Moore, 1983).

    6.1.4.4  Soil and dust

         It is extremely difficult to choose the most appropriate model to
    describe the soil/dust to blood lead relationship, given the many
    variables involved in determining the exposure patterns of children
    and the kinetics involved between the levels in the environment and
    the child. A review of the available studies shows the extreme
    variability in slopes obtained (0.028-0.36 µmol lead/litre
    (0.6-7.6 µg/dl) blood for each 1000 µg/g soil and 0.00096-0.35 µmol
    lead/litre (0.02-7.2 µg/dl) blood for each 1000 µg/g dust consumed by
    children) (US EPA, 1986a). Detailed consideration has been given to
    the process for the assessment of lead contamination in soil and the
    derivation of soil clean-up criteria by Wixon (1991). The guideline
    model uses a PbB target and slope for the soil:blood lead relationship
    in the particular community in order to derive a PbB guideline. For
    house dust a median value of 0.086 µmol lead/litre (1.8 µg/dl) blood
    per 1000 µg lead/g dust in children appears to be based on data of
    reasonable quality, as does the 0.105 µmol lead/litre (2.2 µg/dl)
    blood per 1000 µg lead/g soil from the same authors.

         According to Elwood (1986), studies of the association between
    lead in house dust and PbB have given inconsistent results and, in
    general, the only studies in which statistically significant
    association has been found are those where lead levels in dust are
    quite high. Landrigan et al. (1975) found a highly significant
    association between PbB and dust lead in an area with a mean level of
    4022 µg lead/g dust, but not in two areas with means of 922 and 816 µg
    lead/g dust. The US EPA (1986a) summarized a series of studies from
    which the overall relationship was judged to be 0.86 µmol/litre
    (18 µg/dl) blood per 1000 µg lead/g dust. Duggan (1980) reviewed the
    literature and determined that a slope of 0.24 µmol/litre (5 µg/dl)
    per 1000 µg lead/g dust is reasonable.

         The overall relationship between PbB and dust/soil lead depends
    on the lead concentrations and bioavailability as well as on the
    proximity and linkage between humans and their environment. This
    relationship varies among locales.

    6.1.4.5  Total lead intake

         The non-linear relationship between PbB and total lead intake
    (see section 5.5) is curvilinear across a broad range of PbB values,
    such that the slope decreases with increasing lead levels. A number of
    biological factors may explain the curvilinear relationships, such as
    increased renal clearance with high PbB (Chamberlain, 1983, 1985),

    distributional non-linearities due to differences in lead binding
    sites in different tissues (Hammond et al., 1981; Manton, 1985;
    Marcus, 1985b), or a sizeable pool of mobile lead in bone maintained
    more or less independently of uptake (Rabinowitz et al., 1976;
    Chamberlain, 1983).

    6.2  Distribution

         The initial distribution of lead in the body may depend upon the
    rate of delivery of blood to various organs. However, it would appear
    that distribution occurs in a similar manner regardless of the route
    of absorption (Kehoe, 1987). The distribution of lead in humans under
    environmental exposure conditions reflects the fact that almost all
    exposures are chronic rather than acute.

    6.2.1  Animal studies

         Studies in rats have shown that lead is rapidly distributed into
    soft tissues and subsequently redistributed into soft and mineralizing
    tissues after acute and chronic exposures. After acute inhalation
    (Morgan & Holmes, 1978) or oral (Aungst et al., 1981) exposure, lead
    levels in rats were highest in liver, kidneys and lung, with levels
    increasing in bone as those in soft tissues declined and stabilized.
    Similar distribution patterns were reported for mice exposed for 12
    months to 21.5 µg lead/m3, the highest levels being found in bone
    and the lowest in lung (Keller & Doherty, 1980a).

         Age-related differences in the distribution of lead in
    experimental animals have been reported. Kostial et al. (1978) noted a
    greater retention of lead in suckling rats than in adults; the levels
    in the brain were also higher in the pups. A 2- to 3-fold increase in
    brain lead concentration (highest in hippocampus) was noted after a
    10-fold (0.1 to 1 mg/kg body weight administered by gavage) increase
    in the dose given to 4- to 8-week-old rat pups (Collins et al., 1992).

         Ageing has been shown to alter the pattern of distribution of
    lead in rats administered lead acetate in drinking-water. Juvenile (21
    days old), adult (8 months old) and elderly (16 months old) rats
    received 0, 1.27 and 6.37 mg lead/kg bodyweight for 9.5 months. The
    pattern of distribution, namely femur > liver > brain was similar in
    all age groups. However, age-related increases in brain lead levels
    were noted, along with decreases in femur lead content (Cory-Slechta,
    1990a).

    6.2.2  Human studies

         Lead is distributed to both soft tissues (blood, liver, kidney,
    etc.) and mineralizing systems (bone and teeth). Bone may be affected
    adversely by lead but also serves as the body's major storage site.

    Bone accumulates lead over much of the human life span, and a study of
    the kinetics of distribution is important since bone can, under
    appropriate conditions, pose a risk as a potential endogenous source
    of lead.

         Once absorbed, lead is not distributed homogeneously throughout
    the body but rather into several distinct compartments. Such
    biokinetic movements have been explained by Rabinowitz et al. (1976)
    using a three-compartment model. This model was based on tracer and
    balance data from five healthy men and has been refined by Marcus
    (1985a,b,c). Three pools (blood, bone and soft tissues) were
    identified, with lead having distinct half-lives in each. Blood lead
    was considered the most labile compartment with a half-life of about
    36 days, and bone lead the most stable with a half-life of about 27
    years. Lead in soft tissue had a half-life of approximately 40 days.

         The specific physiological foundations for these biokinetic
    models have received much attention and are currently being refined
    (O'Flaherty, 1991, 1993). A more thorough understanding of bone
    growth, for example, expressed as a series of allometric equations,
    should help improve models, especially when they must be applied to
    growing and maturing humans.

         Under steady-state conditions, about 96% of PbB is in the
    erythrocyte. At PbB concentrations of < 1.92 µmol/litre (40 µg/dl),
    whole blood and serum lead levels increase linearly in a positive
    manner. At higher PbB concentrations a curvilinear relationship is
    apparent and the serum to blood ratio increases dramatically (Manton &
    Cook, 1984). Such kinetic relationships may be altered during
    pregnancy. From  in vitro data (Ong & Lee, 1980), fetal haemoglobin
    appears to have a greater affinity for lead than adult haemoglobin.

         In adults, approximately 94% of the body burden of lead is in the
    bones, whereas only 73% of the body burden in children is located in
    this compartment (Barry, 1975, 1981). In view of the extremely long
    half-life for lead in bone, this compartment can serve as an
    endogenous source of lead to other compartments long after exposure
    ceases (O'Flaherty et al., 1982; Kehoe, 1987). This is due to the
    labile lead compartment in bone (Rabinowitz et al., 1976) and the
    ongoing bone redistribution which is subject to alteration by  in vivo
    metabolic processes (Parfitt, 1990). Although lead in bone generally
    increases continuously with age, there is evidence that lead levels in
    some bones (e.g., mid-femur and pelvic bone) plateau at middle age and
    decrease with further ageing (Drasch et al., 1987). It is difficult,
    however, to determine the role ageing plays in this process compared
    to other biological reactions. Ageing may account in part for the
    20-25% increase in PbB levels in menopausal women noted by Silbergeld
    et al. (1988).

         The metabolism of lead in bone has been summarized in reports by
    Barry (1975, 1981), Drasch et al. (1987), Silbergeld et al. (1988),
    Skerfving (1988) and Rabinowitz (1991).

         Until recently, it was widely held that the human skeletal system
    provided a metabolically inert depository for lead and was of little
    consequence in health-risk assessment. It had been assumed that bone
    lead is metabolically inert, with a half-life long enough to forestall
    the risk of ready transfer back to blood. Current evidence is that
    bone comprises a set of kinetically variable subcompartments for lead
    deposition and is a target for toxicity. These factors complicate bone
    lead kinetics as applied to long-term modelling; the mobility of bone
    lead to blood is important (Rabinowitz, 1991).

         Bone lead is readily mobilized to blood and the effect is most
    apparent in people with a history of occupational exposure; bone lead
    also appears to be a major source of blood lead in older people with
    previous ambient exposures to lead. Of particular importance is
    mobilization of lead from bone in pregnant women and nursing mothers
    (Silbergeld, 1991). The mobilization of lead from bone to the more
    bioavailable maternal blood compartment poses a risk to the fetus and
    mother.

         Human bone appears to have at least two, possibly three,
    kinetically distinct lead compartments. Lead in trabecular (spongy)
    bone appears to be more mobile than lead lodged in cortical (compact)
    bone, and there appears also to be a fraction of bone lead in
    equilibrium with the lead in blood (Skerfving, 1988). Trabecular bone
    seems to be an important source of resorbed lead when high exposure is
    reduced, e.g., through removal of medical reasons by retirement of
    lead workers, or in response to chelation in adults (Shutz et al.,
    1987).

    6.2.3  Transplacental transfer

         Lead is readily transferred from the mother to the developing
    infant during pregnancy and accumulates in bone during gestation
    (Barltrop, 1969). The lead concentration in cord blood is 85-90% that
    of maternal blood. Moore et al. (1982) reported a geometric mean level
    of 0.67 µmol/litre (14 µg/dl) in 236 pregnant women and
    0.58 µmol/litre (12 µg/dl) for lead in umbilical cord blood. The mean
    concentration of lead in umbilical cord blood from a sample of over
    11 000 women was 0.298 µmol/litre (6.6 ± 3.2 µg/dl) (Bellinger et al.,
    1987).

    6.3  Elimination and excretion

         In both humans and experimental animals lead is eliminated from
    the body in both urine and faeces. Any dietary (including waterborne)
    lead not absorbed in the gastrointestinal tract is excreted in faeces.
    Airborne lead that has been swallowed and not absorbed is also

    eliminated in this manner. Blood lead not retained in the body is
    excreted in urine or faeces, the latter by biliary excretion. Adults
    ingesting daily 0.3 to 3.0 mg lead (as lead acetate) in drinking-water
    for 16 to 208 weeks excreted more than 85% of the ingested lead, 90%
    of the excreted lead being in the faeces. The amount excreted through
    any route is affected by age and exposure characteristics and is
    species dependent (US EPA, 1986a). This section reviews only studies
    of lead excretion in humans; a discussion of the results of work on
    experimental animals can be found in reviews by US EPA (1986a) and
    ATSDR (1993).

         The age dependency of lead excretion in humans has not been
    studied extensively. However, the studies of Rabinowitz et al. (1977b)
    and Chamberlain et al. (1978) in adults and Ziegler et al. (1978) in
    infants can be used to assess this phenomenon. Data from Ziegler et
    al. (1978) and Rabinowitz et al. (1977a) are given in Table 20.
    Ziegler et al. (1978) calculated a total daily retention by infants of
    40 µg, which is about twice the amount calculated by Alexander et al.
    (1973).

    Table 20.  Comparison of daily lead intake and excretion in children
               and adultsa
                                                                 

    Parameter                            Childrenb      Adultsc
                                                                 

    Dietary intake (µg/kg)                 10.76         3.63
    Fraction absorbed                       0.55d        0.15
    Dietary lead absorbed (µg/kg)           5.92         0.54
    Air lead absorbed (µg/kg)               0.20         0.21
    Total absorbed lead (µg/kg)             6.12         0.75
    Urinary lead excreted (µg/kg)           1.00         0.47
    Endogenous faecal lead (µg/kg)          1.56         0.24
    Total excreted lead (µg/kg)             2.56         0.71
    Excreted/absorbed lead                  0.42         0.92
    Fraction of intake retained             0.33         0.01
                                                                 

    a  Adapted from: US EPA (1986a)
    b  From: Ziegler et al. (1978)
    c  From: Rabinowitz et al. (1977a)
    d  Corrected for calculated endogenous faecal lead


         Under conditions of relatively constant exposure to low
    concentrations of lead, approximately 140 to 215 µg/day, a steady
    state condition evolves in which excretion approximates intake
    (Rabinowitz et al., 1976). Under these conditions urinary lead
    excretion is approximately 70% of absorbed lead. Chamberlain (1985)
    reported that approximately 60% of absorbed lead is retained by the
    body and 40% excreted.

         Chamberlain (1983, 1985) also examined the relationship between
    the level of exposure and rate of lead excretion. Renal clearance at
    PbB levels of between 1.2 and 3.84 µmol/litre (25 and 80 µg/dl) was
    found to increase at a rate approximating the increase in plasma lead.

         Chamberlain (1985) estimated endogenous faecal lead loss into the
    gastrointestinal tract following administration of lead-203 via the
    inhalation and parenteral routes. These estimates suggested a
    clearance of approximately 0.5% of administered dose per day when PbB
    concentrations were under 1.2 µmol/litre (25 µg/dl). A special form of
    excretion of endogenous lead is through breast milk. Studies of breast
    milk indicate that lead concentrations correlate with maternal PbB
    concentrations, most studies reporting that lead secreted from breast
    milk varies in concentration between 10 and 30% of the maternal PbB
    concentration (Ong et al., 1985).

    6.4  Biological indices of lead exposure and body burden

    6.4.1  Blood lead

         The relationship between levels of exposure from various
    environmental media and PbB has been discussed briefly in section
    6.1.4.

         Due to the ease of sampling and homogeneity of the sample, blood
    has been the most widely used specimen to assess the human body burden
    of lead. However, in view of the relatively short half-life for lead
    in blood (28-36 days) (see section 6.2.2), PbB measurements in general
    reflect only recent exposures. Also, in view of the kinetics of
    distribution within the body (cycling between blood, bone and soft
    tissues), differentiation of low-level chronic exposure from a short
    high-level exposure is not possible on the basis of a single PbB
    measurement. Interpretation of PbB levels over a wide range of values
    must take account of the curvilinear relationship between total intake
    of lead and PbB concentrations, as well as the proportion of lead in
    plasma (Manton & Cook, 1984; see also section 8.1).

         A number of cohort studies have collected serial PbB measurements
    for children from birth up to 7 or 10 years. For children who have not
    had major changes in their environment, there is good correlation
    between consecutive PbB measurements (McMichael et al., 1988;
    Bellinger et al., 1992; Baghurst et al., 1992; Dietrich et al.,
    1993a,b). In these extended studies it has become apparent that for
    most of the children a single PbB analysis at 6 years of age gives a
    reasonable assessment of the life-time lead exposure status of the
    child. However, random PbB levels in samples taken before 6 years of
    age can markedly underestimate the peak exposure usually seen at 2
    years of age (SAHC, 1993).

         A new exposure index, namely "lifetime average blood lead" or
    "lifetime average integrated blood lead" has been introduced in
    studies using serial blood data. It has been clearly explained in the
    proceedings of a recent Workshop and this explanation is quoted
    directly here. This measure, which is a reflection of the area under
    the "blood lead by age curve", has been occasionally misunderstood.
    Due to the shape of the longitudinal blood lead profile, the peak
    blood lead level, usually observed during the second year of life is
    considerably greater than the average lifetime blood lead. For
    example, a five-year-old child with a lifetime average blood lead of
    0.96 µmol/litre (20 µg/dl) is likely to experience blood lead levels
    above 1.92 µmol/litre (40 µg/dl) during the second year of life and
    may spend 3 years of life, from 12 months to 48 months, above
    0.96 µmol/litre (20 µg/dl). Due to the shape of the curve, with
    declining PbB beginning about 24-39 months of age, the average
    lifetime blood lead for a given child decreases with increasing age.
    This does not mean that the index is inappropriate but rather that
    developmental blood lead profiles change over time. For example, a
    lifetime average blood lead of 0.72 µmol/litre (15 µg/dl) should not
    be interpreted as being equivalent to a single blood lead
    determination of 0.72 µmol/litre (15 µg/dl) obtained at a single point
    in life (Dietrich et al., 1991).

    6.4.2  Tooth lead

         Unlike blood samples, teeth are composed of several anatomically
    distinct pools which form over several years. Thus, in contrast to
    blood, teeth are useful tissues for assessing long-term lead
    accumulation from prenatal exposures to the time of shedding of the
    tooth.

         The accumulation of lead in teeth (PbT) has been used as a
    measure of exposure of children to lead in several epidemiological
    studies (Needleman et al., 1972; Winneke et al., 1982a; Rabinowitz et
    al., 1989; Fergusson et al., 1989; Hansen et al., 1989). Studies
    measuring PbT and PbB, however, are few. Preliminary data from
    Australia (Baghurst et al., 1992) indicate a correlation of 0.8
    between whole PbT and PbB prior to tooth exfoliation, but the only
    report relating lead in circumpulpal dentine and longitudinal PbB (by
    Greene et al., 1992) found a correlation of only 0.5. PbT
    concentrations can vary as a function of location of the tooth within
    the mouth, age, and whether total PbT or dentine PbT were reported.
    Therefore, the age of the tooth in the mouth and its location, the
    sample (whole tooth or dentine) analysed must be considered in any
    estimation of a corresponding PbB in the published studies (SAHC,
    1993).

         Because of the complex structure and development of teeth,
    concentrations of PbT will depend on the method of sampling and
    analysis, tooth type, and resorption and tooth age at exfoliation. As

    a result of varying procedures used by different investigators, there
    may be substantial variation in absolute values and possibly the
    biological meaning of PbT levels between different studies.

         Different parts of a tooth sequester lead during different stages
    of development. This is especially of concern in tooth sampling
    because lead is not uniformly distributed within a tooth on a
    submillimeter scale. A pilot study of teeth from children with
    differing histories of exposure to lead and using high precision lead
    isotopic methods has shown that analyses of slices of the incisal part
    of deciduous teeth give the clearest indications of the  in utero
    environment, and the cervical sections, the exposure from birth to
    exfoliation (Gulson & Wilson, 1994).

         For more detailed discussions on the measurement of PbT the
    reader is referred to Grandjean et al. (1984), Purchase & Fergusson
    (1986) and Fergusson et al. (1989).

    6.4.3  Bone lead

         The human skeleton begins to accumulate lead during fetal
    development and continues to abut 60 years of age (Pounds et al.,
    1991). Interest in bone lead and its measurement  in vivo stems from
    concern that skeletal lead is not metabolically inert (see section
    6.2.2), but can be mobilized by physiological and pathological states,
    for example, during pregnancy and lactation (Silbergeld, 1991) and
    osteoporosis (Silbergeld et al., 1988), with possible adverse effects
    in other tissues, including the fetus, as well as the desire to
    develop a meaningful measure of cumulative lead exposure as a tool in
    public health protection. Section 8.13 includes a brief discussion of
    the skeleton as a target organ for lead toxicity.

         Procedures are available to analyse bone samples for lead levels
    in humans not occupationally exposed to lead (Drasch et al., 1987;
    Drasch & Ott, 1988). These studies have shown a decrease in levels of
    lead in bones from autopsy specimens in Germany after removal of lead
    from petrol. However, full utilization of bone lead stores as
    dosimeters of lead exposure in a prospective sense requires the
    utilization of technologies for  in vivo measurement of lead in bone
    such as X-ray fluorescence analysis (Chettle et al., 1991; Rosen et
    al., 1991; Todd et al., 1992).

    6.4.4  Lead in urine

         Although urinary lead level has been used to measure current
    exposure (Robinson, 1974), its use as a biomarker of lead exposure is
    questionable in view of the relatively low and variable level of lead
    excreted in the urine (Jensen, 1984; Ibels & Pollock, 1986). For this
    reason, and in view of technical difficulties in analysing low levels
    of lead in urine, urinary lead appears to be of limited use for

    general screening. However, where an elevated body burden of lead has
    been estimated using PbB or other indices of lead exposure, urinary
    levels of lead after administration of the chelating agent
    CaNa2-EDTA is considered an excellent measure of the potentially
    toxic fraction of the total body burden of lead (CDC, 1985). The
    chelatable lead excreted is assumed to represent lead removal from
    soft tissues and blood, as well as sub-compartments of bone (Ibels &
    Pollock, 1986; Mushak, 1989). In contrast, the urinary lead excretion
    associated with lead mobilization provides what is considered the best
    measure of the potentially toxic fraction of the total body burden
    (see CDC, 1985; US EPA, 1986a). On the basis of various  in vitro
    experimental and epidemiological studies (CDC, 1985; US EPA, 1986a;
    Mushak, 1989), chelatable lead is assumed to be a chemical sample of
    both mobile body compartments (i.e. blood and soft tissues) as well as
    of sub-compartments of bone.

    6.4.5  Lead in hair

         Hair lead has been used as an indicator of exposure in children
    (Marlowe & Errera, 1982; Wilhelm et al., 1989). However, there are
    severe limitations on its use from both the methodological as well as
    the metabolic perspective. Systemic variations in lead level have been
    reported according to hair colour, texture, location on the body and
    growth phase (Wilhelm et al., 1989). Also, it is almost impossible to
    avoid external contamination, and to date no validated methods are
    available for cleaning. Methods which are sufficiently vigorous to
    remove superficial lead also remove lead from the hair shaft.

    7.  EFFECTS ON LABORATORY ANIMALS AND IN VITRO TEST SYSTEMS

         Lead can affect various organ systems depending upon the level
    and duration of exposure. In all species studied, adverse effects on
    the nervous system have been noted at PbB concentrations lower than
    for other target organs. There is particular concern for the effects
    of lead on fetal development. Since the appearance of Environmental
    Health Criteria 3: Lead (IPCS, 1977), there have been very many
    reports on the effects of lead on  in vitro and animal models. No
    attempt will be made in this chapter to summarize all such studies.
    Rather, emphasis has been given to those studies that relate most
    directly to the understanding of the effects of lead on humans and
    thus provide additional scientific support for the use of such human
    studies in the assessment of risk from lead exposures.

    7.1  Biochemical effects

    7.1.1  Haem synthesis and haematopoiesis

         The principal clinical manifestation of the effect of lead on the
    haematopoietic system is anaemia but this occurs only with high levels
    of exposure that are rarely seen today. Lead affects the
    haematopoietic system at several levels. These include effects on haem
    and globin synthesis and on erythrocyte formation and function.

         Most haem synthesis is directed toward formation of haemoglobin
    and the rest is used in cellular oxidative metabolism. Lead inhibits
    several steps in the biosynthesis of haem. More detail can be found in
    IPCS (1977), US EPA (1986a) and section 8.1.1 of this monograph.

         Lead acts on steps in the synthetic pathway both inside and
    outside the mitochondrion. It inhibits certain enzymes (ALA
    dehydratase, ferrochelatase, coproporphyrinogen oxidase) and increases
    the activity of ALA synthase (ALAS) activity as a consequence of
    feed-back regulation by haem (Moore et al., 1980). The activity of
    ALAS is the rate-limiting step in the haem biosynthetic pathway.

         Lead affects erythrocyte formation by impairment of globin and
    haem synthesis. Globin synthesis is inhibited by lead in rat bone
    marrow cells at concentrations as low as 1 µmol/litre (Dresner et al.,
    1982). This is thought to be secondary to decreased protein synthesis
    in erythroid cells as a consequence of lead-induced inhibition of haem
    synthesis. Lead also decreases erythrocyte survival through its
    inhibition of membrane bound Na-K ATPase (Rhagavan et al., 1981).

         It has been shown in  in vitro cultures of cells from liver,
    bone marrow and the nervous system that ALAS activity may be increased
    by addition of only 50 µmol lead/litre and ALAD activity is decreased
    by 60% at a concentration of 0.5 µmol/litre (Kusell et al., 1978;
    Dresner et al., 1982). Lead (5 µmol/litre) also inhibits (20%)
    porphobilinogen deaminase in red blood cell haemolysates at

    5 µmol/litre (Piper & Tephly, 1974). Fowler et al. (1980) have shown
    that disruption of haem synthesis results in reduction of tissue haem
    levels. Increased exposure to lead decreases the content and function
    of haem-dependent enzymes of the P-450 mono-oxygenase system (Meredith
    & Moore, 1979). More recently it has been shown that lead induces haem
    oxygenase activity thereby increasing the degradation of haemoproteins
    which may adversely affect a number of cell functions such as
    respiration and energy production (Maines, 1992). It has been
    suggested that delay in the accumulation of haemoproteins of the
    respiratory chain in brain tissue during development may result in
    decreased synthesis of haem enzymes in the brain (Bull, 1980; Moore et
    al., 1987). Holtzman et al. (1981), on the other hand, found no effect
    of lead on brain cytochromes in rat pups with impairment of growth due
    to exposure to lead.

    7.2  Nervous system effects

    7.2.1  Higher order behavioural toxicity

         Experimental studies using animal models have demonstrated that
    lead impairs learning and memory functions at virtually all stages of
    the life cycle. The most significant studies have focused on
    behavioural and learning impairments in experimental animals with PbB
    levels below 1.44 µmol/litre (30 µg/dl). Bushnell & Levin (1983) fed
    post-weaning rats drinking-water containing 10 or 100 mg lead (as lead
    acetate) per litre for up to 7 weeks. PbB was not determined but has
    been estimated from comparisons with similar lead exposure models to
    have been less than 0.96 µmol/litre (20 µg/dl) (Davis et al., 1990b).
    Brain lead averaged about 0.05 µg/g wet weight. The rats exhibited
    impaired learning ability when tested for their ability to choose
    between alternate arms of a radial maze. Cory-Slechta et al. (1985)
    demonstrated a significant learning impairment in rats given lead
    acetate in drinking-water (25 mg/litre) from weaning. The observed
    outcome was a significantly higher response rate in lead-exposed rats
    working under a fixed interval schedule of food reinforcement. PbB
    values of lead-treated rats in that study averaged
    0.72-0.96 µmol/litre (15-20 µg/dl) and brain lead levels averaged
    0.07 µg/g wet weight. Similar effects have been described in older
    rats (16 months of age) at steady-state PbB levels of
    0.62-0.86 µmol/litre (13-18 µg/dl) (Cory-Slechta & Pokora, 1991). More
    recently, Cohn et al. (1993) demonstrated selective effects of lead on
    learning processes, as distinct from non-specific effects such as
    motivational levels, and sensory or motor impairment using a multiple
    schedule of repeated learning and performance. Exposure to lead
    acetate in drinking-water (50 mg/litre) produced a PbB of
    1.2 µmol/litre (25 µg/dl).

         Winneke et al. (1977) exposed rats to lead while  in utero,
    through mother's milk and directly in drinking-water. Their learning
    ability was then tested by requiring the animals to discriminate
    between stimuli of either different orientation (stripes as an easy

    task) or different size (discs as a difficult task). Adult rats tested
    between 90 and 170 days of age with PbB levels of less than
    1.44 µmol/litre (< 30 µg/dl) were slower to learn the difficult task
    of size discrimination (but not the easy discrimination problem) and
    tended to repeat more errors than control subjects.

         Altmann et al. (1993) demonstrated deficits of both active
    avoidance learning (AAL) and  in vitro hippocampal long-term
    potentiation (LTP) in adult rats that had received pre-weaning or pre-
    and post-weaning dietary lead exposure to achieve PbB levels of
    approximately 0.72 µmol/litre (15 µg/dl) and brain lead concentrations
    of 0.09-0.16 µg/g wet weight. Deficit in animals with only
    post-weaning exposure and blood lead levels was about 0.77 µmol/litre
    (16 µg/dl) but brain lead concentrations of only 0.09 µg/g was either
    absent (in the case of LTP) or markedly reduced (in the case of AAL).

         Similar types of effects have been noted in studies using
    non-human primates. For example, Rice (1985) reported deficits of
    discrimination reversal performance in monkeys dosed orally with 0, 50
    or 100 µg lead/kg body weight per day for the first 200 days of life.
    At the cessation of dosing, PbB concentrations were 0.144 µmol/litre
    (3 µg/dl), 0.72 µmol/litre (15 µg/dl) and 1.2 µmol/litre (25 µg/dl),
    respectively. Prior to behavioural testing (3 years of age), PbB
    levels were 0.144 µmol/litre (3 µg/dl), 0.528 µmol/litre (11 µg/dl)
    and 0.624 µmol/litre (13 µg/dl). Additional evidence for lead-induced
    changes in learning in non-human primates is provided by the study of
    Lilienthal et al. (1990), which showed a dose-related increase in the
    percentage of errors in a learning set formation task (discrimination)
    with PbB levels in the low exposure group (350 µg lead acetate/g diet)
    averaging 1.68 µmol/litre (35 µg/dl). Also, an impairment of reversal
    learning was found by Bushnell & Bowman (1979b) to persist in monkeys
    up to their fifth year, at which time the PbB level in treated animals
    averaged 0.24 µmol/litre (5 µg/dl) compared to 0.192 µmol/litre
    (4 µg/dl) in controls. In the study by Bushnell & Bowman (1979a), the
    period of lead exposure was limited to the first year of life and
    resulted in PbB levels averaging as high as 3.12 µmol/litre
    (65 µg/dl). No description was given of quality control procedures to
    ensure accuracy of PbB determinations at the lower levels reported.

         Several studies on experimental animals have shown perseverative
    effects (the tendency to respond repetitively and inappropriately even
    though environmental conditions have changed) which may underlie many
    of the lead-induced changes in learning and other higher order
    behavioural processes noted in such studies. This has been clearly
    demonstrated in rodents by Cohn et al. (1993) in the context of
    learning impairments. In the study by Bushnell & Levin (1983)
    described above, the accuracy impairment derived from a tendency of
    lead-exposed rats to re-enter a previously explored arm of the maze.
    Similar perseverative tendencies have been described in non-human 

    primate studies, e.g., on delayed matching to sample (Rice, 1984a) and
    delayed alteration tasks (Levin & Bowman, 1986; Rice & Karpinski,
    1988).

         Experimental animal studies also reveal the importance of task
    complexity in detecting lead-induced changes in behaviour, both in
    non-human primates (Levin & Bowman, 1983, 1986; Gilbert & Rice, 1987)
    and rodents (Winneke et al., 1977, 1982b).

         Persistence of lead-induced changes in some higher order
    behavioural processes is also suggested by a number of experimental
    animal studies. Several reports from the longitudinal studies of
    non-human primates by Bowman and his colleagues (e.g., Bushnell &
    Bowman, 1979a,b; Levin & Bowman, 1983, 1986, 1989) have demonstrated
    persistent neurobehavioural deficits extending up to 8 years after the
    termination of exposure and long after PbB levels had declined to the
    control level. In rats, Cory-Slechta & Thompson (1979) showed that the
    levels of lead exposure determine the persistence of lead effects,
    while Cory-Slechta (1990a,b) revealed that apparently transient
    effects of lead on behaviour could re-emerge with changes in the
    reward contingencies of the environment. Munoz et al. (1986) found
    deficits in both spatial and visual discrimination performance of rats
    at several months of age resulting from pre-weaning exposure via the
    dams. The same outcome was noted by Altmann et al. (1993) in a study
    where deficits in adult rats were noted following pre-weaning
    exposure.

    7.2.2  Mechanisms of lead-induced behavioural toxicity

         While the mechanisms underlying lead-induced behavioural toxicity
    have yet to be adequately determined, the experimental animal
    literature provides suggestive leads. The early experimental studies
    of Pentschew (1965) and Pentschew & Garro (1966) demonstrated in the
    rodent model that the pathogenesis of acute high-dose encephalopathy
    was secondary to increased permeability of capillaries in the brain,
    leading to leakage of fluid and red blood cells. These changes are
    similar to those occurring in children with acute lead encephalopathy
    characterized clinically by coma, convulsions and death, and identify
    the brain microvasculature as the primary target for lead in the
    central nervous system following high level exposure to lead.

         Changes in microvascular morphology are not evident with low
    level lead exposure, but it is the vulnerability of the blood-brain
    barrier that permits exposure of the brain to lead. Exposure of the
    developing fetus to lead results in higher uptake of lead in the brain
    than from later exposures (Rossouw et al., 1987). The development of
    resistance to acute encephalopathy in rat pups during the first few
    days of life is thought by Holtzman et al. (1984) to be related to
    maturation of the blood-brain barrier and possibly to the ability of

    the older animals to sequester lead in protein complexes. Thus
    neuropathological changes may relate more to the issue of exposure
    than to mechanism of effect.

         More recently, the focus of mechanistic studies has involved
    biochemical and neurochemical changes. Biochemical changes in synaptic
    transmission are likely to be related to problems in dendrite-nerve
    organization and function (Goldstein, 1990). There is continuing
    reorganization of dendrite and nerve terminal connections throughout
    early months of neural development. A number of biochemical and
    functional studies concerning possible mechanisms of lead effect
    suggest lead may disrupt this process and its function through early
    childhood (e.g., Cookman et al., 1987, 1988).

         Other biochemical effects from lead exposure which may be the
    basis of important neurobiological mechanisms of leadinduced
    behavioural toxicity include changes in protein kinases, with
    implications for neurotransmitter system disturbances. At least three
    protein kinases present in nerve terminals involved in modulating the
    release of neurotransmitters have been shown to be affected by lead.
    However, the relevance of these findings to cognitive function has not
    been established (Goldstein, 1990).

         From both  in vivo and  in vitro studies, lead exposure has
    been reported to have effects on virtually all neurotransmitter
    systems (US EPA, 1986a). Depending on the stage of development, these
    include dopaminergic, cholinergic, serotonergic, GABAnergic,
    glutamatergic and opiate systems. Probably the most extensively
    investigated system has been the dopaminergic system where a number of
    changes at the biochemical and receptor level have been described. One
    of the difficulties, however, is determining the relationship of the
    reported changes at the biochemical/receptor level to changes in
    behavioural function. McIntosh et al. (1989) noted lead-induced
    changes in tetrahydrobiopterin metabolism which may be related to
    changes in IQ scores (Blair et al., 1982; McIntosh et al., 1985). In
    further support of a functional role for lead-induced changes in
    dopaminergic systems, Cory-Slechta & Widzowski (1991) reported
    alterations in both D1 and D2 dopaminergic sensitivity in rats with
    PbB levels of 1.2-1.44 µmol/litre (25-30 µg/dl).

         More recently, studies by Altmann et al. (1993) and Cohn &
    Cory-Slechta (1993) suggest a role for changes in the N-methyl-D-
    aspartate (NMDA) receptor complex in lead-induced behavioural
    toxicity. The NMDA receptor complex has been implicated in the
    learning and memory processes.

         Effects of lead on mitochondrial energy metabolism may also be
    important in the pathogenesis of neurological effects. Impairment of
    respiration in brain mitochondria has been observed in  in vitro
    preparations (Holtzman et al., 1978) and in mitochondria isolated from
    brains of lead-exposed rats (Gmerek et al., 1981).

    7.2.2.1  Conclusions

         Given their ability to establish causal relationships between
    lead exposure and biological effects, experimental animal studies can
    provide evidence supportive of human epidemiological findings. Many
    experimental animal studies have been carried out to characterize the
    nature of lead's effects on various target organ systems as well as to
    establish the underlying mechanisms of effect. Furthermore, these
    studies, by their very nature, are not confounded by such co-variates
    of children's IQ as parental IQ, socioeconomic status and quality of
    the home environment, or complicated by nutritional inadequacies of
    the study population. Earlier experimental animal studies tended to
    utilize relatively high lead exposure levels, but over the past ten
    years, exposure levels and protocols have been more relevant to the
    human situation.

         There are, of course, some limitations of experimental animal
    studies that must be considered. For example, there may be differences
    in species sensitivity as well as differences in pharmacokinetic and
    pharmacodynamic behaviour of lead in rodents, primates and humans.
    Higher external exposure levels of lead are required to produce PbB
    levels corresponding to levels experienced in human populations, and
    higher PbBs are necessary to induce encephalopathy in rodents than are
    associated with such effects in humans. Such differences in
    sensitivity obviously reflect, at least in part, differences in
    kinetics of lead across species. Despite these differences, it should
    be noted that corresponding levels of PbB have been associated with
    neuro-behavioural toxicity in both animal models (rodents as well as
    non-human primates) and human populations. This comparability of
    effect levels, despite the relative insensitivity of the animal
    models, suggests that the levels of concern in humans might be even
    lower, although quantitative extrapolation is difficult. Although one
    cannot presume that the neurotoxic effects of lead in experimental
    animals necessarily predict specific effects in humans, the
    similarities in neurobehavioural end-points in humans and animals are
    sufficient to conclude that neurobehavioural deficits in animals are
    at least qualitatively predictive of effects in humans.

    7.2.3  Sensory organ toxicity

         Visual and auditory functions have been shown to be affected by
    lead; other sensory modalities have largely been neglected. Bushnell
    et al. (1977b) studied visual acuity in infant rhesus monkeys using
    different degrees of illumination. Four control animals and three
    animals each in a low and a high lead group were tested. The animals
    were orally dosed between postnatal days 5 and 365. Although the
    dosing regimen of this group was intended to produce blood lead levels
    around 4.08 µmol/litre (85 µg/dl), peak levels of between 6.58 and
    14.4 µmol/litre (137 and 300 µg/dl) occurred between the fifth and the
    ninth weeks of dosing. Significant impairment of scotopic vision was
    observed only in the animals of the high lead group.

         Using flash-evoked visual potentials (VEP), Lilienthal et al.
    (1986) studied visual function in 7- to 7´-year-old rhesus monkeys
    pre- and postnatally exposed to lead acetate (0, 350 and 600 mg/kg) in
    their diets. Seventeen animals were tested altogether: six controls,
    and five or six in each of the exposed groups. PbB levels fluctuated
    around 1.92 µmol/litre (40 µg/dl) in the group fed 350 mg/kg diet
    during the seven postnatal years, whereas in the group fed 600 mg/kg
    they had declined from the initial average of 5.28 µmol/litre
    (110 µg/dl) at age 9 months to about 2.88 µmol/litre (60 µg/dl) at 7
    years of age. PbB levels in the mothers of both these groups had been
    between 1.15 and 1.78 µmol/litre (24 to 37 µg/dl) during pregnancy. A
    dose-related decrease of amplitudes and a similar increase of
    latencies for the main VEP component were found; these were, in most
    instances, significant in both the high and low lead groups.

         These findings are supported by the results of studies on rodents
    by Fox et al. (1977). Prolonged VEP latencies were observed following
    chronic developmental lead exposure. Lower VEP amplitudes have also
    been observed in rats at PbB levels exceeding 1.44 µmol/litre
    (30 µg/dl) (Winneke, 1979).  in vitro studies suggest that rods are
    more sensitive to the effects of lead than are cones (Fox & Sillman,
    1979).

         Information on the impairment of auditory function by lead comes
    from electrophysiological studies using brainstem auditory evoked
    potentials (BAEP) in rhesus monkeys (Lilienthal et al., 1990). PbB
    levels in the various groups at the time of testing averaged
    0.432 µmol/litre (9 µg/dl; controls), 1.92 µmol/litre (40 µg/dl; low
    lead), and 2.688 µmol/litre (56 µg/dl; high lead). Of the four peaks
    discernible in the monkey, BAEP peaks II and IV exhibited
    significantly prolonged latencies, which were observed only in the
    animals of the high lead group.

    7.3  Renal system

         The renal effects of lead in animal models occur as a result of
    both acute and chronic exposures. Acute lead nephrotoxicity is
    characterized by decreased reabsorption of small molecular weight
    compounds by the renal tubule, particularly amino acids, glucose and
    phosphate (Fanconi Syndrome).

         Morphological changes of chronic lead exposure include
    cytomegaly, development of nuclear inclusion bodies and
    ultrastructural changes in mitochondria. The basis for the cytomegaly
    is not understood but it is known that there is altered homoeostasis
    of water and electrolytes and cellular swelling. Inclusion bodies are
    lead-protein complexes composed of acidic non-histone proteins. As
    much as 90% of lead in the kidney has been shown to be contained in
    the inclusion bodies, suggesting that they provide a detoxification
    function. The origin of the protein is not known but Egle & Shelton
    (1986) have identified the most abundant component of isolated

    inclusion bodies to be a constitutive protein of the adult central
    nervous system primarily in the cerebral cortex. The lead in the
    inclusion bodies is chelatable with EDTA (Goyer & Wilson, 1975).

         Mitochondria isolated from kidneys of lead-intoxicated rats have
    impaired respiration and oxidative phosphorylation capacity, (Goyer &
    Rhyne, 1973).

         Kholil-Manesh et al. (1992) studied the evaluations of renal
    histological and functional changes in rats continuously dosed with
    high dose lead. Glomerular filtration rate (GFR) was significantly
    increased at 3 months, but significantly decreased at 12 months. Lead
    inclusion bodies were noted throughout the duration of the study.
    Tubular atrophy and intestinal fibrosis appeared at 6 months. The
    brush border of proximal tubular cells was disrupted at 1 and 3
    months, but recovered later.

         With continued exposure to lead, acute nephropathy may progress
    to chronic interstitial nephritis that does not have any unique or
    distinguishing features. There is progressive increase in interstitial
    fibrosis, dilatation of tubules and formation of microcysts with
    hyperplasia of tubular epithelial cells. In the rat, inclusion bodies
    are reduced in number and may be entirely absent in later stages of
    the nephropathy. Glomerular sclerosis occurs with onset of proteinuria
    and renal failure (Goyer & Rhyne, 1973).

         The sequence of morphological changes observed in experimental
    models is thought to be generally true for humans. Experimental
    studies in rats suggest that there may be a threshold for lead renal
    toxicity in the rat at PbB levels similar to those observed for
    humans. In rats exposed to different doses of lead for up to 12 weeks,
    a PbB level of 2.88 µmol/litre (60 µg/dl) appears to be the threshold
    for proximal renal tubular cell injury by lead (Goyer et al., 1989).
    This PbB level is equivalent to kidney lead levels of about 45 mg/kg
    wet weight and is the level of lead at which excretion of renal
    calcium increases and ultrastructural changes occur in proximal
    tubular cell mitochondria. This finding is consistent with the
    observation in humans by Buchet et al. (1980) and Gennart et al.
    (1992) of a threshold value of 2.88 µmol/litre (60 µg/dl) of PbB which
    leads to adverse renal effects by lead.

    7.4  Cardiovascular system

         Early studies concerning the production of hypertension in
    experimental animals, summarized in Environmental Health Criteria 3:
    Lead (IPCS, 1977), were conflicting. Among rats given 70 mg of lead
    acetate per day orally, only a few survived 40 days and all were
    hypertensive (Griffith & Landaver, 1944). On the other hand, other
    studies did not reveal a blood pressure effect from high-level lead
    exposure in rats (Padilla et al., 1969) or in dogs (Fouts & Page,

    1942). More recently, several experimental studies have confirmed that
    lead can produce hypertension, and evidence for several plausible
    mechanisms has been provided. Rats, both normotensive and
    spontaneously hypertensive, exposed to lead in water supplies for up
    to a year, resulting in PbB levels of up to 0.6 µmol/litre
    (12.5 µg/dl), exhibited ventricular tachycardia and ventricular
    fibrillation (Evis et al., 1985). Subsequent studies (Evis et al.,
    1987) showed that adrenaline had a significant arrhythmogenic effect
    in hypertensive animals. It was concluded that chronic exposure to
    lead, when combined with high blood pressure, slightly enhances the
    susceptibility of the heart to arrhythmias induced by myocardial
    ischaemia. Overviews of the experimental evidence related to
    hypertension and lead exposure have been published by Victery (1988)
    and US EPA (1986a, 1989). Taken as a whole, these studies demonstrate
    that increase in blood pressure does occur secondary to renal failure
    in rats with continuous high-level exposure to lead. More recent
    studies on effects of chronic low-level exposure to lead on blood
    pressure in rats have shown alterations in cardiovascular parameters
    in the PbB range of 0.24-1.92 µmol/litre (5-40 µg/dl), and provide
    some insights as to mechanisms involved in the pathogenesis of
    lead-related cardiovascular effects.

         Victery et al. (1982a) exposed rats to lead  in utero by giving
    dams drinking-water containing 0, 5, and 25 µg lead/ml and continued
    this regimen to the pups for 5 to 6 months. Although no change in
    blood pressure was noted at these levels of exposure, significant
    changes in the renin-angiotensin system were reported in animals given
    water containing 25 µg lead/ml with a PbB level of 0.864 µmol/litre
    (18 µg/dl). In a similar study (Victery et al., 1982b), exposure to
    100 µg lead/ml in drinking-water (but not 500 µg/ml) produced a
    significant (17 mmHg) elevation in blood pressure beginning at 3´
    months of age and continuing until 6 months of age. Chai & Webb (1988)
    found an elevation of 15 to 20 mmHg in the systolic blood pressure of
    rats given drinking-water containing 100 µg lead/ml. These authors
    suggest that alterations in the cellular mechanisms that regulate
    intracellular calcium concentration may enhance pressor responsiveness
    to catacholamines. Boscolo & Carmignani (1988) found raised blood
    pressure in rats with exposure to drinking-water containing 0, 30 or
    60 µg lead/ml for 18 months. Cardiovascular responses to blood
    pressure agonists indicated that lead exposure affects the
    renin-angiotensin system and induces sympathetic hyperactivity by
    acting on central and peripheral sympathetic junctions, increasing the
    responsiveness to stimulation of alpha-2-adrenergic receptors, and by
    increasing sensitivity to stimulation of cardiac and vascular
    beta-adrenergic and dopaminergic receptors.

         From the experimental evidence summarized by Victery (1988), it
    appears that low-level exposure to lead produces an elevation in blood
    pressure. The failure to demonstrate increased blood pressure in some
    studies with high-level exposure to lead suggests that the effect of

    lead on blood pressure may be biphasic, i.e. a consistent effect with
    low-level exposure but inconsistent effects with high-level exposure.

    7.5  Reproductive system

         Experimental studies on the effects of lead on the reproductive
    system most often concern toxicity to either the male or female but
    have addressed results of exposure to both parents. Environmental
    Health Criteria 3: Lead (IPCS, 1977) identified effects on
    spermatogenesis in rats exposed to lead and showed that high maternal
    exposure to lead in rats can reduce numbers and size of offspring.
    There may also be paternally transmitted effects resulting in
    reductions of litter size, weights of offspring and in survival rate.

         Few studies into the effects of lead on male sexual function have
    reported PbB levels. Ivanova-Chemishanska et al. (1980) reported
    changes in levels of enzymatic activity and ATP in testicular
    homogenates from rats given 0.0001 or 0.01% solutions of lead acetate
    as drinking-water over a 4-month period. Chowdhury et al. (1984) found
    testicular atrophy along with cellular degeneration in rats with PbB
    levels over 3.36 µmol/litre (70 µg/dl), but not in rats with levels of
    2.59 µmol/litre (54 µg/dl). Donovan et al. (1980) found that lead
    inhibited androgen binding by cytosolic receptors in mouse prostate.
    Testicular homogenates from 2- to 3-week-old male offspring (PbB
    levels 0.30 µmol/litre, 6.3 µg/dl) of lead-exposed female rats showed
    a decreased ability to metabolize progesterone (Wiebe et al., 1982).
    In an  in vitro study, Wiebe et al. (1983) found a 10 to 20% decrease
    in FSH binding and in the production of cyclic AMP by Sertoli cells
    isolated from prepubertal rats and cultured in the presence of lead
    acetate (2.64 × 10-4 mol/litre). In addition, the activity of
    cellular 3ß-hydroxysteroid dehydrogenase was decreased. It was shown
    by Sokol et al. (1985) that there is a dose-related suppression of
    serum testosterone levels and spermatogens in adult rats (100 day of
    age) given a solution of 0.3% sodium acetate as drinking-water for up
    to 60 days. PbB concentrations were between 1.44 and 2.40 µmol/litre
    (30 and 50 µg/dl), depending on the length of treatment. Further
    studies (Sokol, 1987) supported the hypothesis that lead disrupts the
    hypothalamic control of pituitary hormone secretion. However, other
    evidence indicates that lead may directly or indirectly affect
    testicular enzymes or may act indirectly by a reduction in testicular
    binding of FSH and production of cyclic AMP (US EPA, 1986a).

         Dosing mature female rats with lead in order to produce PbB
    concentrations of 1.44 µmol/litre (30 µg/dl) resulted in irregular
    estrous cycles. At a PbB level of 2.54 µmol/litre (53 µg/dl), animals
    developed follicular cysts and there was a reduction in the number of
    corpora lutea (Hilderbrand et al., 1973). Grant et al. (1980) reported
    delayed vaginal opening in rats whose mothers were given drinking-
    water containing 25, 50 and 250 µg lead/ml. The vaginal opening delays
    in the group given 25 µg lead/ml of drinking-water occurred in the

    absence of any growth retardation or other developmental delays and
    were associated with median PbB levels of 0.86-1.39 µmol/litre
    (18-29 µg/dl). Studies on female monkeys have shown that pre- and/or
    postnatal exposure to lead can affect pubertal progression and
    hypothalamicpituitary-ovarian-uterine functions. Chronic exposure of
    nulliparous female monkeys to lead (PbB levels of approximately
    1.68 µmol/litre, 35 µg/dl) resulted in subclinical suppression of
    circulating luteinizing and follicle stimulating hormone and estradiol
    without producing overt effects on general health or menstruation
    (Foster, 1992).

    7.6  Effects on bone

         There is growing interest in lead in bone for several reasons.
    Bone is a store for lead accumulated from past exposure. It has a long
    biological half-life but may be mobilized and contribute to blood lead
    during pregnancy. With the development of X-ray fluorescence
    techniques to measure lead in bone  in vivo, there is a need to
    improve our understanding of bone lead metabolism and of factors that
    influence lead retention and release. In addition, lead may adversely
    affect bone metabolism, particularly in post-menopausal women, and
    contribute to the development of osteoporosis. There has only been
    limited experimental study of these concerns to date, but there are
    some relevant reports in the literature.

         A summary of much of the currently available literature on the
    potential toxicological implications of lead in bone during pregnancy
    and lactation is contained in a review by Silbergeld (1991). Toward
    the latter half of pregnancy in mice, there seems to be a preferential
    transfer of lead across the placenta to the fetus (Danielson et al.,
    1983). In rats exposed to lead for 150 days and then not exposed for
    50 days prior to mating, Buchet et al. (1977) found that there was a
    substantial mobilization of lead from mother to fetus. Keller &
    Doherty (1980b) found using radiotracer lead (210Pb) in female mice
    that there was also a major transfer of lead to the pup during
    lactation. The concomitant decrease in maternal bone lead supports the
    hypothesis that bone resorption of lead occurs during lactation.

         It has been shown that lead may directly and indirectly affect
    various aspects of bone metabolism (Pounds et al., 1991). Lead
    inhibits the renal enzyme 1-hydroxylase, reducing plasma levels of
    1,2-dihydroxychole-calciferol (activated vitamin D). Lead impairs the
    Haversian remodelling system in beagle dogs chronically exposed to
    lead (Anderson & Danylchuk, 1977).

    7.7  Immunological effects

         The effects of lead on the immune system are diverse but not well
    documented. Lead reduces resistance and increases mortality of
    experimental animals when they are infected by a broad range of

    bacterial and viral agents (Koller, 1984). Lead impairs antibody
    production in animals and generally decreases immunoglobin plaque-
    forming cells (Koller & Roan, 1980).

    7.8  Mutagenicity

         Lead is thought to have genotoxic properties. However, lead-
    induced gene mutations in cultures of mammalian cells have only been
    observed at concentrations toxic to the cells. Studies for point
    mutations in bacterial systems have also yielded negative results (US
    EPA, 1986a). Zelikoff et al. (1988) found that both insoluble lead
    sulfide and more soluble lead nitrate were mutagenic when added to
    Chinese hamster V79 cells. A 6-fold increase in mutation frequency was
    noted at a lead nitrate level of 500 µmol/litre medium. These authors
    also found that lead acetate induced morphological transformation of
    Syrian hamster cells. However, they concluded that these effects may
    not have been the result of direct damage to DNA but may have occurred
    via indirect mechanisms including disturbances in enzyme functions
    important in DNA synthesis and/or repair.

         Studies on the production of chromosome aberrations, sister-
    chromatid exchanges and micronuclei by lead, whether in  in vitro
    cultures or  in vivo, have given mixed results, and summaries are
    available (US EPA, 1986a; IARC, 1987a,b; ATSDR, 1991).

    7.9  Carcinogenicity

         There have been several experimental studies in rats and mice in
    which long-term administration of a lead compound in food or
    drinking-water or parental administration has produced tumours of the
    kidney (Van Esch & Kroes, 1969; Moore & Meredith, 1979). These and
    other studies have been discussed in detail in IARC (1980) and
    summarized in US EPA (1986a). In general, methodological problems,
    including dose levels, number of animals, and doses, and lack of
    toxicity monitoring, make many of the studies difficult to interpret
    in a quantitative manner. One study (Azar et al., 1973) addressed some
    of these problems and reported an increase in the numbers of renal
    tumours in rats fed diets containing 500 mg lead/kg for 2 years. PbB
    levels were 3.84 µmol/litre (80 µg/dl). In all studies renal
    carcinogenicity occurred against a background of proximal tubular cell
    hyperplasia, cytomegaly and cellular dysplasia in response to the high
    doses of lead and long exposure times (Goyer, 1985). Renal
    adenocarcinoma occurred in a high percentage of lead-exposed animals
    and the incidence of tumours was related to the length and severity of
    exposure (Mao & Molnar, 1967). Males appear to be more susceptible to
    tumours than females. The maximum dose of lead in drinking-water, not
    associated with any morphological or functional evidence of renal
    toxicity in rats fed a diet containing adequate levels of trace
    minerals, particularly calcium, is 200 mg lead/litre (Goyer et al.,
    1970). No evidence of renal tumours at doses below this value has been
    reported (Azar et al., 1973; US NCI, 1979).

         Several hypotheses have been proposed for the mechanism of lead
    carcinogenicity in experimental animals. These include mutagenicity,
    cellular proliferation, nuclear protein (inclusion bodies), promoter
    activity, activation of protein kinase C, and cystic hyperplasia
    (Goyer, 1993). Lead is a weak mutagen in mammalian cell systems, but
    is a strong mitogen. Exposure to a single intraperitoneal injection of
    lead acetate in rats stimulates a 40-fold increase in cell
    proliferation as measured by autoradiography, and this is further
    increased by unilateral nephrectomy (Choie & Richter, 1972). DNA
    synthesis in kidneys, as measured by 3H-thymidine incorporation, is
    increased 15-fold and the mitotic index 45-fold following a single
    intracardiac injection of lead acetate in mice (Choie & Richter,
    1974). Cell proliferation and hyperplasia are seen in the liver of
    rats given a single intravenous injection of lead nitrate (Columbano
    et al., 1984). In both of these studies, increases in cell
    proliferation occurred in the absence of cellular necrosis, suggesting
    that this was a mitogenic rather than regenerative response.

         Activation of protein kinase C and formation of nuclear inclusion
    bodies or lead protein complexes are additional events that may
    influence regulation of cell growth and development and play a role in
    the carcinogenic response in experimental animals resulting from
    administration of lead (Goyer, 1993). Cystic hyperplasia, a late
    morphological manifestation of chronic lead nephropathy, is a risk
    factor for renal cancer (Bernstein et al., 1987). Prior to adenoma
    formation in animals treated with renal carcinogens, cystic
    hyperplasia was reported (Dees et al., 1980; Goyer et al., 1981).

    8.  EFFECTS ON HUMANS

         Despite the long recognition of lead poisoning, new clinical
    cases continue to be the subject of published reports. Although
    important, they affect a comparatively small proportion of the
    population at risk from the potential effects of exposure to
    environmental lead.

         Over the last 10-15 years, particular attention has been directed
    towards epidemiological studies designed to evaluate the possible
    neurotoxic effects of lead on the developing child, especially delayed
    or impaired neurobehavioural development and performance.

         In the adult population, considerable attention has been directed
    towards evaluating cardiovascular effects and the implication of lead
    in hypertension.

         New studies have contributed to our understanding of the
    biochemical effects of lead, and may facilitate early recognition of
    significant change and mitigation of potentially adverse outcomes.

         The effects of lead have been studied widely in both the general
    population and in those exposed occupationally. Since these effects
    are the same in both settings, no distinction has generally been made
    in the discussion. However, it is often important to distinguish
    between adults and children because of different susceptibility.

    8.1  Biochemical effects of lead

         Lead is known to affect a number of enzymes and physiological
    systems which result in a wide variety of changes in humans. While
    those affecting the haematopoietic system are well known, there are
    others which need to be considered in the risk assessment process and
    are considered in this section.

         In considering the effects of lead on biochemical systems, it is
    appropriate to discuss the form of the lead in the various body
    compartments.

         Blood lead (PbB) is distributed between the plasma and the
    erythrocyte. There is less than 1% in the plasma for PbB levels of up
    to 4.8 µmol/litre (100 µg/dl) (Manton & Cook, 1984). The curvilinear
    relationship of serum lead to blood lead is shown in Fig. 12. The data
    show that the erythrocytes have a capacity to bind lead up to PbB
    levels of about 2.4 µmol/litre (50 µg/dl). Above this level a fairly
    rapid increase in the serum levels occurs.

         It should be noted that the proportion of "free" (i.e. unbound)
    lead in blood is important in relation to biological activity. Lead is
    bound to haemoglobin in blood and has a greater affinity for fetal
    than adult haemoglobin (Ong & Lee, 1980). It may, therefore, be

    FIGURE 12

    important to consider the proportion of fetal haemoglobin present in
    blood samples from mothers and infants in assessing PbB concentrations
    in relation to biological effects. Also, increased fetal haemoglobin
    has been found in cases of human and experimental animal poisoning
    (Albahary, 1972).

    8.1.1  Haem synthesis

         Lead is known to affect several enzymatic reactions critical in
    haem synthesis, causing abnormal concentrations of haem precursors in
    blood and urine. These effects of lead on haem synthesis are shown in
    Fig. 13.

         As shown in Fig. 13, lead inhibits the activity of three enzymes
    of the biosynthetic pathway, 5-aminolaevulinate dehydratase (ALA-D),
    coproporphyrinogen oxidase (COPRO-O) and ferrochelatase (FERRO-C).
    This depletes haem synthesis and depresses the synthesis of the
    initial and rate-limiting enzyme 5-aminolaevulinate (ALA) synthase. As
    a consequence there is increased production and excretion of the
    precursors ALA and coproporphyrin (COPRO) with increased circulatory
    protoporphyrin (PROTO) usually bound to zinc. In the red cell,
    diminished synthesis of monooxygenases (cytochromes P450) compromises
    drug oxidation and lead is bound to haemoglobin.

    8.1.1.1  Protoporphyrin levels

         Lead interferes with the conversion of protoporphyrin to haem by
    ferrochelatase. The protoporphyrin exists under these circumstances
    primarily as zinc protoporphyrin, with a proportion remaining free
    (Chisolm & Brown, 1979).

         The relationship between protoporphyrin, either free or as the
    zinc chelate, in blood and blood lead is one which could be
    interpreted as showing a "threshold of effect" or as a continuum of
    effect. The exact mathematical relationship is inevitably the choice
    of the investigator since the uncertainties in the measurement of
    blood lead and blood protoporphyrin would allow the fitting of the
    data to either of these relationships. The threshold concept is
    satisfactory to those that seek the "no-effect level" (Succop et al.,
    1989), since it provides the start point of an analytical
    identification process. The continuum of effect is, however, much more
    plausible in a biological sense, since the idea of a level at which
    "no effect" will occur is unlikely in a biological system.

         The figures for the relationship between PbB and zinc
    protoporphyrin are confused in any population by iron status (Marcus &
    Schwartz, 1987), and it is thus unlikely that measurement of zinc
    protoporphyrin, either alone or compared with haemoglobin status, will
    provide an accurate estimation of PbB levels less than 0.96 µmol/litre
    (20 µg/dl). In a study of the relationship between PbB levels and
    erythrocyte protoporphyrin concentration for 2004 urban children

    FIGURE 13

    (Piomelli et al., 1982), it was concluded that a threshold is apparent
    and occurs at a PbB level of between 0.72 and 0.86 µmol/litre (15 and
    18 µg/dl). This is consistent with the data of Roels et al. (1976),
    which show a clear discontinuity at around 1.2 µmol/litre (25 µg/dl).
    Marcus & Schwartz (1987) re-analysed data from the NHANES II survey of
    264 children. No positive correlation between PbB and zinc
    protoporphyrin was noted below 0.96 µmol/litre (20 µg/dl).

         Other reports which present relevant data are Roels et al.
    (1976); Piomelli et al. (1982); Hammond et al. (1985); Rabinowitz et
    al. (1986); Roels & Lauwerys (1987). In the presence of iron
    deficiency the observed threshold is likely to be lower (Mahaffey &
    Annest 1986; Marcus & Schwartz, 1987). At first sight it would seem
    inappropriate to consider thresholds that are determined in the
    presence of potential iron deficiency.

         A study by Koren et al. (1990) of maternal and umbilical cord
    lead and free erythrocyte porphyrin (FEP) levels for 95 mother-infant
    pairs showed a correlation between maternal and cord PbB, with
    maternal levels exceeding neonatal levels. Most cord PbB levels were
    below 0.33 µmol/litre (7 µg/dl), and 11 were below the detection
    limit. The cord blood FEP (0.86 µmol/litre) level was consistently
    higher than maternal FEP (0.53 µmol/litre) but was not statistically
    correlated. The elevated cord blood FEP values were attributable to
    immature haem synthesis and high erythrocyte volume rather than the
    presence of lead.

    8.1.1.2  Coproporphyrin levels

         One of the earliest observed effects of lead poisoning was a rise
    in coproporphyrin excretion in the urine, due to inhibition of
    coproporphyrinogen oxidase (Campbell et al., 1977). Although often
    cited as a good measure of lead exposure it is clear that the
    importance of the measure of current environmental levels of exposure
    is small. This is because the excretion levels do not rise
    significantly until the PbB excretion is greater than 1.92 µmol/litre
    (40 µg/dl) (Meredith et al., 1978).

    8.1.1.3  delta-Aminolaevulinic acid levels in urine and blood

         Like protoporphyrin, circulating and excreted levels of ALA are
    likely to be best described as a continuum of effect. Elevated levels
    of this compound are of importance since neurological features of lead
    exposure have been ascribed in part to increased circulating levels of
    ALA (Moore et al., 1987). The rise in concentration during lead
    exposure is a function first of decreased activity of ALA dehydratase
    (ALAD), which is uniquely sensitive to lead toxicity, and subsequently
    of increased activity of the initial and rate-limiting enzyme of haem
    biosynthesis, ALA synthase (Meredith et al., 1978). It is
    inappropriate to discuss haem biosynthesis control mechanisms here;

    suffice it to say that blocking this pathway by lead lowers free haem
    levels, which feedback to ALA synthase (Moore et al., 1987). The
    special relationship between ALAD activity and lead exposure has been
    best described as a negative exponential and has been used as a
    measure of lead exposure in population surveys (Berlin & Schaller,
    1974).

         The immediate effect of the inhibition of ALAD will be an
    increased level of ALA in the blood, which will then lead to increased
    urinary excretion. The plasma levels of ALA are elevated in the
    presence of higher lead levels. This has been seen in a number of
    studies (Haeger-Aronsen, 1960; Meredith et al., 1978; O'Flaherty et
    al., 1980).

         Meredith et al. (1978) measured ALA metabolism in 48 male
    lead-exposed workers (aged 22-56 years, PbB 4.2 ± 1.4 µmol/litre,
    87 ± 29 µg/dl), who were compared with control subjects (28 male, 9
    female, 18-52 years of age, PbB 1.3 ± 0.4 µmol/litre, 27 ± 3.3 µg/dl).
    They found increasing levels of circulating ALA associated with PbB
    that reached a plateau when the PbB level was in excess of
    3 µmol/litre (62 µg/dl) and ALA exceeded 4 µmol/litre. At higher blood
    ALA levels, urinary excretion of ALA increased exponentially,
    consistent with decreased tubular reabsorption. The authors suggested
    that there was a "critical" tissue lead concentration of around
    2 µmol/litre (41 µg/dl). This study showed some continuity of the
    correlation down to the lowest PbB value of the control group, namely
    0.86 µmol/litre (18 µg/dl). However, these data were interpreted to
    show that effects are only demonstrable above a PbB level of
    1.44 µmol/litre (30 µg/dl).

         Other studies show direct correlations between PbB level and
    urinary ALA (Selander & Cramer, 1970; Lauwerys et al., 1974), although
    these correlations are not seen as low as those in the study of
    Meredith et al. (1978). The data of Selander & Cramer (1970) showed a
    clear threshold effect at about 1.02 µmol/litre (40 µg/dl) in
    occupational subjects.

         Roels et al. (1976) reported data over a range of PbB levels from
    0.24 to 1.92 µmol/litre (5 to 40 µg/dl) in children which showed
    essentially no correlation with urinary ALA.

         Data obtained from 39 men and 36 women in the general population
    showed that increased urinary excretion of ALA occurred at PbB levels
    of more than 1.68 µmol/litre (35 µg/dl) in women and more than
    2.16 µmol/litre (45 µg/dl) in men (Roels & Lauwerys, 1987). The
    sensitivity of the haem synthesis pathway to increased lead exposure
    was in the order: children > women > men.

         On balance, it would appear that lead has discernible effects on
    the urine level of ALA at a PbB level of around 1.68 µmol/litre
    (35 µg/dl).

    8.1.1.4  delta-Aminolaevulinic acid dehydratase levels

         delta-Aminolaevulinic acid dehydratase (ALAD) is commonly
    measured for its activity in samples of haemolysed blood; the result
    may or may not be corrected for haematocrit value. The measure of
    enzyme activity will reflect the amount of enzyme present as well as
    the effect of inhibition of the enzyme.

         Studies in the general population have confirmed the correlation
    and the apparent lack of a threshold for inhibition of ALAD in
    different age groups and exposure categories (Roels et al., 1976;
    Chisolm et al., 1985; Roels & Lauwerys, 1987). A negative linear
    relationship between PbB and ALAD activity was found between mothers
    and their newborn babies (cord blood); PbB levels ranged from 0.14 to
    1.44 µmol/litre (3-30 µg/dl) (Roels et al., 1976). Roels & Lauwerys
    (1987) reported a similar relationship in a population of 143 children
    aged 10-13 years having PbB levels of 0.19-1.97 µmol/litre
    (4.7-41 µg/dl).

         In the study of Roels et al. (1976) a large number of children
    were studied. Although the authors drew a regression line which
    suggested that the effects continued to very low levels, the data
    between 0.24-0.72 µmol/litre (5-15 µg/dl) were very scattered, and the
    regression at low levels appeared to be rather speculative. It was
    considered that the lead level above which an effect level is
    demonstrable from these data was 0.48 µmol/litre (10 µg/dl).

    8.1.1.5  delta-Aminolaevulinic acid synthase

         Aminolaevulinic acid synthase (ALAS) levels are determined by a
    feedback mechanism which is dependent on haem levels. There appear to
    have been relatively few studies on serum ALAS, although some
    consideration is given in a report by Meredith et al. (1978). However,
    insufficient data were available to establish effect levels for lead
    exposure.

    8.1.1.6  Other effects of decreased haem synthesis

         The potential impact of a reduction in the body pool of haem and
    haem precursors is shown in Fig. 14. It is evident that the multiple
    effects on haem metabolism from lead exposure can lead to adverse
    effects in organs and systems other than the erythropoietic system.

    8.1.2  Vitamin D

         Formation of the most important vitamin D metabolite,
    1,25-dihydroxyvitamin D, is by 1 alpha-hydroxylation of
    25-hydroxyvitamin D in the kidney. This is mediated by
    25-hydroxyvitamin D-1 alpla-hydroxylase, a cytochrome P450-dependent
    enzyme in the mitochondria of the renal tubules. Serum concentrations

    of 1,25-dihydroxyvitamin D are measured in children as an indicator of
    the effects of lead on the enzyme system mediating the initial
    hydroxylation. However, other factors such as dietary intake and the
    physiological needs for calcium and phosphorus, and levels of
    calciotropic hormones such as parathyroid hormone, can regulate the
    production and circulating concentrations of 1,25-dihydroxy-vitamin D
    (Rosen & Chesney, 1983).

         Several studies have provided information on the effect of lead
    on the circulating concentrations of 1,25-dihydroxyvitamin D. Rosen et
    al. (1980) studied children with PbB levels in the range of
    1.58-5.76 µmol/litre (33-120 µg/dl). While the most striking decreases
    in serum 1,25-dihydroxyvitamin D occurred in children whose PbB level
    was 2.97 µmol/litre (62 µg/dl), the effect was considered to be
    evident in the range of 1.58-2.64 µmol/litre (33-55 µg/dl) when
    compared to an age- and race-matched control group with PbB levels in
    the range of 0.48-1.248 µmol/litre (10-26 µg/dl).

         Mahaffey et al. (1982) measured serum 1,25-dihydroxyvitamin D in
    177 subjects aged 1-16 years. PbB lead measurements were performed in
    105 of these children and ranged from 0.576 to 5.76 µmol/litre (12 to
    120 µg/dl). A curvilinear relationship between serum
    1,25-dihydroxyvitamin D concentration and PbB was noted in children
    aged 2 to 3 years. Details of the dietary intake of the subjects were
    not available.

         Koo et al. (1991) studied 105 children, aged 1 to 3 years and
    with detailed lead exposure history from birth, to determine the
    effect of chronic low to moderate lead exposure. The average lifetime
    PbB concentration was 0.23-1.13 µmol/litre (4.8-23.6 µg/dl) and was
    greater than 0.96 µmol/litre (> 20 µg/dl) in only three children.
    With a range of concurrent PbB concentrations of 0.29-2.11 µmol/litre
    (6-44 µg/dl). The children generally had adequate dietary intakes of
    calcium, phosphorus and vitamin D. An effect of lead was found on
    serum concentrations of calcium, phosphorus, 25-hydroxyvitamin D,
    1,25-dihydroxyvitamin D and parathyroid hormone, and on bone mineral
    content. In the presence of adequate nutritional status, lead exposure
    at low levels (leading to a PbB level of less than 0.96 µmol/litre or
    20 µg/dl) appears to have no demonstrable effect on circulating
    concentrations of total and ionized calcium, magnesium, phosphorus,
    calciotropic hormones including 1,25-dihydroxyvitamin D, parathyroid
    hormone and calcitonin, or bone mineralization as indicated by single
    photon absorptiometry. At higher levels there is a demonstrable effect
    on 1,25-dihydroxyvitamin D.

    8.1.3  Dihydrobiopterin reductase

         This enzyme (DHBR) is part of the synthesis/salvage cycle which
    controls the hydroxylation of tyrosine, tryptophan and phenylalanine.
    It is thus central to the synthesis of the catecholamines. Lead has
    been shown to inhibit the synthesis of tetrahydrobiopterin from
    dihydrobiopterin by dihydrobiopterin reductase (Leeming & Blair, 1980;

    FIGURE 14

    McIntosh et al., 1985) in rat brain. In humans the exact relationship
    between PbB and plasma biopterins has not been properly defined.

    8.1.4  Nicotinamide adenine dinucleotide synthetase

         A recent study of nicotinamide adenine dinucleotide (NAD)
    synthetase in erythrocytes indicates that this enzyme is sensitive to
    inhibition by lead (and zinc) and is a sensitive indicator of lead
    exposure. Zerez et al. (1990) found NAD synthetase activity reduced in
    three subjects with elevated PbB levels in the range
    1.63-3.46 µmol/litre (34-72 µg/dl).

    8.1.5  Nutritionally affected groups

         A body of evidence, summarized and discussed by Mahaffey (1985),
    indicates that a high dietary calcium intake tends to decrease lead
    absorption and retention in infants, young children and adults. Other
    evidence suggests that in groups including children, with
    self-selected diets, low dietary calcium intakes are associated with a
    greater prevalence of elevated PbB levels (Mahaffey, 1985).

         Some human data indicate that adults with low iron states have an
    increased absorption of both iron and lead (Watson et al., 1980,
    1986). In children, particularly those from low-income families, there
    is often an association between iron deficiency and elevated PbB (Yip
    & Dallman, 1984; Mahaffey, 1985). While this by itself does not
    indicate a causal relationship, when considered in the context of data
    on adults and animals, it suggests that iron deficiency in children
    can result in increased lead absorption.

         Alcoholics and people who consume excess alcohol may be at
    increased risk of adverse effects from lead. In animal studies,
    alcohol and lead synergistically inhibit ALAD, hepatic glutamic
    oxaloacetic transaminase and glutamic pyruvic transaminase activities
    (Flora & Tandon, 1987) supporting the hypothesis regarding risks to
    humans.

    8.2  Haematopoietic system

         Lead-induced anaemia can be a direct consequence of inhibition of
    haem biosynthesis; it is not necessarily associated with iron
    deficiency. It may be associated with alterations of globin syntheses
    (Albahary, 1972). More importantly, the synthesis of alpha- and
    ß-globin chains may become asynchronous (White & Harvey, 1972).

    8.2.1  Anaemia

         Based on data published by Lilis et al. (1978), Baker et al.
    (1979), Grandjean (1979), and earlier data, the threshold PbB level
    for a decrease in haemoglobin has been estimated to be 2.40 µmol/litre
    (50 µg/dl) for occupationally exposed adults (US EPA, 1986a).

         Grandjean et al. (1989a) demonstrated the reduced ability of the
    erythropoietic system to regenerate after blood withdrawal
    (0.45 litre) in 25 lead-exposed battery workers (average PbB level
    2.14 µmol/litre, 44.5 µg/dl) compared with 25 age-matched controls
    (average PbB level 0.35 µmol/litre, 7 µg/dl). The haematological
    parameters, except for erythropoietin were otherwise comparable for
    the two groups. The effect was attributed to an effect on haem
    synthesis and possible decreased erythrocyte survival in lead-exposed
    workers. Lead has been found to depress serum levels of
    erythropoietin, a hormone which regulates erythrocyte formation
    (Graziano et al., 1991), which might also affect the reserve capacity
    for erythropoiesis.

         The PbB threshold for decreased haemoglobin levels in children is
    estimated to be approximately 1.92 µmol/litre (40 µg/dl) (IPCS, 1977).
    However, a cross-sectional epidemiological study of 579 children, aged
    1-5 years in 1974, living in close proximity to a primary lead smelter
    showed that adverse effects on haematocrit may occur at lower PbB
    levels (Schwartz et al., 1990). Anaemia, defined as a haematocrit
    below 35%, was not found at PbB levels of less than 0.92 µmol/litre
    (20 µg/dl). There was a strong non-linear dose-response relationship
    at higher PbB levels, which was influenced by age.

         Kutbi et al. (1989) studied 200 boys aged 6-8 years. The mean PbB
    was 0.33 µmol/litre (6.9 µg/dl; range 0.07-1.14 µmol/litre or
    1.4-23.8 µg/dl). Subdividing the group at 0.72 µmol/litre (15 µg/dl),
    they found a negative correlation between PbB level and all
    haematological values for the "upper normal" (N = 7) compared with the
    "normal" group (N = 193). The pattern of haematological parameters was
    described as predictive of microcytic anaemia.

         Anaemia has been commonly associated with the adverse effects of
    occupational lead exposures. It is an effect that is easily diagnosed
    clinically and is recognized as a marker of lead toxicity. Anaemia may
    result from either a decrease in haemoglobin production or an increase
    in the rate of destruction of erythrocytes. An analysis, made in 1974,
    of the association between PbB levels and haematocrit in 579 children
    (1-5 years of age) living near a primary lead smelter has recently
    been presented (Schwartz et al., 1990). A haematocrit value of less
    than 35% was used to indicate an adverse effect. It should be noted
    that the effect of iron deficiency was not taken into account in
    analysing the results. The study concluded that there was no adverse
    effect of lead at PbB levels below 0.96 µmol/litre (20 µg/dl).
    Furthermore, the risk of having a haematocrit value below 35% for
    1-year-olds was 2% at PbB levels between 0.96 and 1.87 µmol/litre
    (20 and 39 µg/dl). The degree of iron deficiency may account for a
    substantial proportion of this 2%. In this study, the level at which
    an effect of lead on the induction of anaemia was demonstrable was
    about 1.92 µmol/litre (40 µg/dl).

    8.2.2  Pyrimidine-5'-nucleotidase activity

         Inhibition of erythrocyte pyrimidine-5'-nucleotidase leads to
    accumulation of pyrimidine nucleotides, which has been associated with
    induction of basophilic stippling.

         Impaired function of erythrocyte pyrimidine-5'-nucleotidase
    occurs in lead-exposed workers and as a genetically induced enzyme
    deficiency. In  in vitro systems the enzyme is inhibited by lead,
    cadmium, mercury and zinc ions (and perhaps other metal ions) (Paglia
    et al., 1975; Pagliuca et al., 1990; Ichiba & Tomokuni, 1990).

         Graphs of PbB concentration against erythrocyte pyrimidine-
    5'-nucleotidase activity in children have been reported for 21
    children aged 2-5 years (Angle & McIntire, 1978) and for 42 aged 1-5
    years (Angle et al., 1982) who participated in a preventative
    programme because they had previously had PbB levels above
    1.44 µmol/litre (30 µg/dl) or zinc protoporphyrin (ZPP) greater than
    60 µg/dl. Both plots show considerable scatter although there was a
    significant inverse logarithmic correlation. However, data relating to
    the reproducibility, precision and accuracy of assessing inhibition of
    erythrocyte pyrimidine-5'-nucleotidase were not presented.

         This test may be capable of analytical refinement when assessment
    of the sensitivity to lead (and other metals) exposure might more
    reasonably be made. It does not yet appear likely to provide a routine
    measure for low-lead exposure assessment.

    8.2.3  Erythropoietin production

         Graziano et al. (1991) found depressed serum erythropoeitin
    levels in females at mid-pregnancy and at delivery associated
    statistically with PbB level. Erythropoietin is a glycoprotein
    produced in the renal proximal tubules which regulates both
    steady-state and accelerated erythrocyte production. The study was not
    adequate to define a dose-response relationship.

    8.3  Nervous system

    8.3.1  Historical perspective

         Gross toxic effects of lead on the nervous system were reported
    by ancient Greek physicians. A brief summary of such reports, and
    those of Roman physicians and 18-19th century toxicologists, was
    published by Kazantzis (1989). The syndrome was known as "painter's
    colic", which included abdominal pain, constipation and paralysis,
    symptoms now covered by the term "lead encephalopathy".

         Lead colic was usually accompanied by effects on the nervous
    system (lead palsy) in cases of acute and chronic poisoning, which
    were often but not always fatal. Lead poisoning was well recognized as

    an occupational hazard. However, it was also associated with
    consumption of lead-contaminated water, wine, cider, rum from the West
    Indies, and food prepared or stored in lead or lead-glazed utensils.
    For example, Kazantzis (1989) described cases of lead poisoning in
    girls resulting from consumption of drinking-water kept in lead-lined
    cisterns, and also described the Devonshire colic attributed by Sir
    George Baker in 1767 to the consumption of cider prepared in presses
    made of lead or lead alloys.

         The early toxicologists clearly recognized that the dose
    (exposure) and form of lead was important in relation to absorption
    and were able to demonstrate the presence of lead in tissues of fatal
    cases of poisoning. There are many individually described, detailed
    case studies in their writings.

    8.3.2  Neurotoxic effects in adults

         Despite the long recognition of lead poisoning, new clinical
    cases continue to be the subject of published reports (Cueto et al.,
    1989; Kocak et al., 1989; Zuckerman et al., 1989; Friedman &
    Weinberger, 1990; Gupta et al., 1990; Mitchell-Heggs et al., 1990;
    Nosal & Wilhelm, 1990; Sharma et al., 1990; Schneitzer et al., 1990;
    Veerula & Noah, 1990). Although they are important, the numbers of
    these poisoning cases are comparatively small compared with the
    population at risk to the potential effects of exposure to lead from
    environmental and dietary sources. Of particular recent concern have
    been the possible neurotoxic effects on the developing child.

    8.3.2.1  Central nervous system

         In a study of 158 secondary lead smelter workers, Lilis et al.
    (1977) found that 64% reported CNS symptoms and that there was early
    occurrence of symptoms within 1 year of exposure. PbB concentrations
    were elevated (about 3.36 µmol/litre or 70 µg/dl) at the time of
    examination but many of the subjects had received chelation therapy.

         Fischbein et al. (1979) noted CNS symptoms in 21 of 81 employees
    of law enforcement agencies working in firing ranges. Symptoms
    correlated with PbB levels and were present in three-quarters of
    subjects having a PbB level > 2.4 µmol/litre (50 µg/dl).

         Hanninen et al. (1979) reported that 49 lead-exposed workers
    (including 21 male and 8 female storage battery and 16 railroad
    engineering workshop employees) whose PbB level had never exceeded
    3.36 µmol/litre (70 µg/dl) experienced more subjective symptoms than a
    control group; excess symptoms were proportional to lead uptake.

         Fischbein et al. (1980) studied 90 telephone cable splicers
    having intermittent exposure to lead. They reported that 26 complained
    of CNS symptoms (mean PbB level of 1.36 µmol/litre, 28.4 µg/dl, and
    ZPP of 70.3 µg/dl) compared with controls (mean PbB level of
    1.32 µmol/litre, 27.4 µg/dl, and ZPP of 49.3 µg/dl).

         Awad El Karim et al. (1986) studied 92 lead acid battery workers
    and compared them with 40 oil mill worker controls. They found CNS
    symptoms in 50% of lead-exposed workers and measured PbB levels in 46
    subjects working in different sections of the plant. More than 95% of
    exposed workers had PbB levels above 1.92 µmol/litre (40 µg/dl),
    whereas the mean PbB in the control group was 1.0 µmol/litre
    (21 µg/dl).

         Impairment of psychological and neuropsychological test
    performance has been reported for lead-exposed workers (Hanninen et
    al., 1979; Hogstedt et al., 1983; Parkinson et al., 1986; Araki et
    al., 1986a,b; Huang et al., 1988a; Stollery et al., 1989, 1991).

         The critical flicker fusion test has been reported to be a
    sensitive indicator of CNS changes associated with exposure to lead.
    Wooller & Melamed (1978) found significantly lower critical flicker
    fusion in subjects with PbB levels above 2.88 µmol/litre (60 µg/dl)
    than in those with levels below 0.96 µmol/litre (20 µg/dl). Williamson
    & Teo (1986) found a significant decrease in mean flicker fusion in 59
    lead-exposed workers (mean PbB > 2.3 µmol/litre or 48 µg/dl) compared
    with matched controls. However, Gennart et al. (1992b) found no
    difference between lead-exposed workers (mean PbB level
    2.45 µmol/litre or 51 µg/dl, N = 98) and controls (mean PbB level
    1.00 µmol/litre or 20.9 µg/dl, N = 85).

         Stollery et al. (1989) studied the performance of 91 men,
    occupationally exposed to inorganic lead, in a series of tests
    designed to assess cognitive function. Subjects were grouped in "low"
    (mean PbB 0.48 µmol/litre or 10 µg/dl, mean ZPP 8.7 µg/dl, N = 28),
    "medium" (mean PbB 1.44 µmol/litre or 30.1 µg/dl, mean ZPP 22.4 µg/dl,
    N = 27) and "high" (mean PbB 2.43 µmol/litre or 50.7 µg/dl, mean ZPP
    58.8 µg/dl, N = 36) ranges of PbB concentrations. Results showed that
    occupational exposure to lead impaired performance in a range of tests
    (sensory motor reaction time, memory, attention, verbal reasoning,
    spatial processing). Workers having PbB levels in excess of
    1.92 µmol/litre (40 µg/dl) showed clear evidence of impairment on
    tests of serial reaction time and category search. The authors
    concluded that sensory motor, rather than cognitive, requirements of
    many psychological tasks provide the most sensitive index of the early
    effects of chronic low level lead exposure.

         In a follow-up study (Stollery et al., 1991), 70 workers were
    re-tested (three times within 8 months). Performance deficits in the
    "high" (mean PbB 2.49 µmol/litre or 51.8 µg/dl, mean ZPP 77.4 µg/dl,

    N = 22) group were not altered by practice or continued exposure. The
    main deficit was a slowing of sensory motor reaction time coupled with
    difficulties in remembering incidental information. There was little
    evidence of impairment in workers having PbB levels less than
    1.92 µmol/litre (40 µg/dl).

         Finally, it appears that neuroelectrophysiological testing is a
    sensitive and objective indicator of the CNS effects of lead. The
    results of Araki et al. (1986b, 1987, 1992) using short-latency
    somatosensory and visual evoked potentials and auditory event-related
    potential (P300) indicate that subclinical electrophysiological
    effects of lead occur not only in peripheral nerves but also in the
    CNS.

    8.3.2.2  Peripheral nervous system

         Peripheral neuropathy is a common sign of chronic, high level
    lead exposure, often manifesting as weakness in the upper or lower
    limbs. At lower levels of lead exposure, nerve conduction velocity
    (NCV) has provided a more sensitive indicator of peripheral nerve
    dysfunction. More than 30 published studies have measured the
    conduction velocity of electrically stimulated sensory and motor
    nerves in workers exposed to lead. However, these studies have yielded
    somewhat mixed results, with many showing a decrease in NCV in
    relation to lead exposure (generally indexed as PbB concentration) and
    a few showing no effect or occasionally even an increase in NCV
    associated with lead exposure (Seppäläinen & Hernberg, 1980; Davis &
    Svendsgaard, 1990).

         Various reasons may underlie this lack of uniformity in NCV
    findings. For example, studies may differ in methodological features,
    in the characterization of lead exposure (e.g., single time-point
    versus time-weighted average PbB levels) or in the handling of
    confounding variables such as nerve temperature and age of subject
    (Ehle, 1986). Other important factors accounting for some of the
    apparent inconsistency in this area of research may be the possible
    antagonistic effect of zinc to lead and differences in nerves selected
    for measurement in different studies (e.g., slow versus fast fibres
    (Araki et al., 1986c; Murata & Araki, 1991). A statistical
    meta-analysis and critical review of 32 NCV studies by Davis &
    Svendsgaard (1990) indicated that the median motor nerve shows effects
    of lead more reliably than other nerves (e.g., median sensory or
    ulnar).

         Despite these complications, certain key well-conducted studies
    provide compelling evidence of a causal relationship between lead
    exposure and reduction in NCV. Araki et al. (1980) measured median
    motor NCV before and after PbB levels of workers were lowered through
    chelation therapy. Depending on the amount of reduction in PbB level
    achieved by chelation and on a given worker's baseline NCV,

    significantly improved NCV was measured in 7 out of 14 lead-exposed
    workers. For all 14 workers the improvement in NCV was significantly
    correlated with the decrease in PbB level (r = -0.573, P < 0.001).

         In another key study, Seppäläinen et al. (1983) followed workers
    prospectively from the beginning of their employment in a battery
    plant, dividing them into two exposure categories (above and below the
    median PbB level of 1.44 µmol/litre (30 µg/dl). Median motor NCV was
    significantly reduced in the workers with blood lead levels above
    1.44 µmol/litre (30 µg/dl) at 1-, 2-, and 4-year evaluation points,
    despite attrition (and reduced statistical power) that resulted in
    only five subjects per group after 4 years. Based on this well-
    conducted prospective study, 1.44 µmol/litre (30 µg/dl) would appear
    to be the lowest-observed-adverse-effect level for reduced NCV in
    adults. Although another well-conducted prospective study (Spivey et
    al., 1980) was unable to find an effect of lead on NCV at higher
    exposure levels, NCV in the median nerve was not measured in that
    study. Triebig et al. (1984) reported a dose-dependent decrease in NCV
    at PbB levels above 3.36 µmol/litre, no dose-effect relationship being
    detectable below this concentration.

    8.3.2.3  Autonomic nervous system

         There have been two reports of tests for autonomic nervous
    function following lead exposure. Teruya et al. (1991) conducted a
    cross-sectional survey of 172 male workers exposed to lead (mean PbB
    level, 1.73 µmol/litre or 36 µg/dl, range 0.24-3.648 µmol/litre or
    5-76 µg/dl), by measuring electrocardiographic R-R interval
    variability (CVRR). Age-adjusted CVRR during deep breathing in
    workers with PbB levels above 1.44 µmol/litre (> 30 µg/dl) was
    significantly decreased as compared to those with PbB levels less than
    0.96 µmol/litre (< 20 µg/dl). In addition, significant dose-response
    and dose-effect relationships were observed between the PbB levels and
    R-R interval variation during deep breathing among the workers with
    PbB levels above 0.96 µmol/litre (> 20 µg/dl).

         In a study reported by Murata & Araki (1991), the R-R interval
    (CVRR) and two component coefficients of variation in the R-R
    interval (C-CVRSA) were measured in 16 gun metal foundry workers
    exposed to lead, zinc and copper (mean PbB level, 1.63 µmol/litre or
    34 µg/dl, range 0.768-2.88 µmol/litre or 16-60 µg/dl). The CVRR and
    component C-CVRSA (respiratory sinus arrhythmia) were significantly
    lower in the workers than in age-matched controls, whereas the
    C-CVMWSA (Mayer wave-related sinus arrhythmia) was unaffected. The
    authors noted that zinc may have been antagonistic to the lead effects
    in this study. Since the CVRR during deep breathing and the
    component C-CVRSA of respiratory sinus arrhythmia reflect
    parasympathetic activity, these two reports indicate potential
    dysfunction of the autonomic nervous system (mainly, parasympathetic
    activity) at average PbB levels of approximately 1.68 µmol/litre
    (35 µg/dl).

    8.3.3  Neurotoxic effects in children

         The majority of the epidemiological research on the health
    effects of lead has been focused on children because, in comparison
    with adults, they are more vulnerable to lead in several respects
    (Davis & Grant, 1992). For example, children typically engage in
    hand-to-mouth activities (sucking fingers, putting food or other
    objects in the mouth) that result in greater ingestion of lead than
    adults normally experience. Also, because of their greater absorption
    and retention of lead, the body burdens in children resulting from a
    given external exposure level tend to be higher than adults.
    Furthermore, the relatively greater exposures and body burdens of
    children occur during sensitive periods of development. Finally, it
    appears that children are generally more sensitive to the
    toxicological effects of lead at a given internal exposure (PbB)
    level. The lowest-observed-effect levels for various end-points (e.g.,
    encephalopathy, anaemia, reduced haemoglobin, elevated EP, slowed NCV,
    impaired neurobehavioural function) are lower in children than in
    adults (US EPA, 1986a, 1990).

         Section 8.3.3.2 describes findings from the main epidemiological
    cross-sectional studies that have been published since 1979. This
    constitutes a very substantial collection of research findings but has
    certain interpretive limitations, particularly the lack of information
    on exposure history.

         More recent research effort has concentrated on prospective
    epidemiological studies of birth cohorts repeatedly evaluated, mostly
    up to the school age years. These prospective studies are described in
    section 8.3.3.3.

         Section 8.3.3.4 presents a quantitative overview of the
    collective evidence from the prospective studies and discusses their
    interpretation and the problems of inferring causality from such
    observation data.

    8.3.3.1  Historical perspective

         While there have undoubtedly been cases of lead poisoning in
    children as long as lead has been used by man, most of the attention
    was previously given to cases of poisoning as a result of occupational
    exposure in adults. In 1892 Gibson and colleagues in Australia
    reported a case series of ten young children with lead colic. It was
    not until twelve years later that peeling paint in the children's
    homes was identified as the lead source (Gibson, 1904).

         In 1943, Byers & Lord (1953) reported the results of an important
    study, the first to suggest that there were neurological after-effects
    of lead poisoning, in the absence of cerebral oedema and high
    intracranial pressure, following acute lead encephalopathy. However,
    in other cases it was felt that neurological manifestations

    disappeared if the ingestion of lead was stopped (McKhann & Vogt,
    1933). The case series of 20 school-age children had been hospitalized
    for lead poisoning but without signs of acute lead encephalopathy, and
    all were later discharged as cured. At follow-up, only one of the
    children was making satisfactory progress at school. The children were
    described as showing a variety of symptoms, including poor academic
    achievement, intellectual deficits, sensory-motor deficits and
    behaviour disturbance.

         In the years following the findings of Byers & Lord (1953), a
    number of case control studies were carried out that examined mental
    retardation (Beattie et al., 1975) or hyperactivity (David et al.,
    1972).

         In addition, as a result of the growing concern about the dangers
    of lead encephalopathy, studies were set up to investigate the
    possibility that subclinical levels of lead (i.e. levels producing no
    overt signs of lead encephalopathy) cause more subtle neurological
    damage to children. Many of these studies investigated children
    identified by screening clinics as having raised lead burdens, but
    others were sited around smelters or lead works where children were
    found to have elevated lead levels. The studies were important in
    determining the areas of functioning which might be affected, and in
    drawing attention to the methodological problems involved in carrying
    out such studies. The main difference between the smelter studies or
    studies around lead works, and clinical studies is that in the former
    the primary source of environmental lead is known. In the hospital or
    clinical studies it is likely that the ingestion of paint flakes, or
    less often leaded putty, was in most cases the primary cause of the
    excess lead exposure, although the exact reason for the increased lead
    burden is rarely known.

         Clinical and smelter studies were carried out in the United
    Kingdom and USA and in general investigated children with PbB levels
    above 1.92 µmol/litre (40 µg/dl), and compared their performance with
    that of children with levels below this value.

         Although some of the clinical and smelter studies found a
    lead-associated deficit (Perino & Ernhart, 1974; Landrigan et al.,
    1975), others did not (Lansdown et al., 1974). In general the studies
    were methodologically flawed, with small samples lacking in
    statistical power, biased ascertainment of subjects or controls, and
    inadequate control for potentially confounding co-factors. The large
    majority did not control for parental intelligence, which is the
    variable most strongly associated with child IQ. These studies were,
    however, important in suggesting an association between body lead
    levels and performance, and also in demonstrating that there is an
    association between social disadvantage and higher body lead levels.

    8.3.4  Population-based cross-sectional studies on children

         Most of the general population-based epidemiological
    cross-sectional studies have, in addition to intelligence tests,
    included extensive batteries of tests covering several other areas of
    psychometric functioning, such as academic attainment, behaviour,
    gross motor abilities and fine motor co-ordination, reaction time,
    visuo-spatial skills and memory, and auditory memory. There are
    differences in the tests used, and in the areas of functioning
    investigated, and the results are less consistent and more difficult
    to interpret than those relating to IQ tests. For that reason this
    section will focus mainly on results relating to intelligence tests,
    and results from other functional domains will be mentioned briefly
    where appropriate.

         There are many cross-sectional population studies using either
    blood or teeth as the primary measure of lead body burden, and only
    the more informative, in terms of providing information on risk
    assessment, will be included. The majority of the studies described
    used either a full or short form of the Wechsler Intelligence Scale
    for Children - Revised (WISC-R or its translations) as the measure of
    intelligence. Exceptions are the studies of Fulton et al. (1987),
    which used the British Ability Scales (BAS), Harvey et al. (1988),
    which used Wechsler Pre-school and Primary Scale of Intelligence
    (WPPSI), and Schroeder et al. (1985) and Hawk et al. (1986), where the
    Stanford-Binet Intelligence Scale was used.

    8.3.4.1  Tooth lead studies

    a)   Needleman et al. (1979)

         A total population of 3329 children aged 6-7 years, attending
    schools in two working class towns near Boston, USA, were asked to
    donate their shed deciduous teeth. Of these 70% (2335) children did
    provide teeth. A 1-mm slice of tooth consisting primarily of dentine,
    was split in half, and one portion was analysed by ASV. Children whose
    initial tooth slice was in the highest 10th percentile (> 20 µg/g) or
    the lowest 10th percentile (< 10 µg/g) were provisionally classified
    into a high or low lead group. A total of 524 children were originally
    classified into these lead groups, but 254 of these children were not
    tested for a number of reasons, and the results for another 112 were
    excluded from data analysis for medical reasons, or because a second
    tooth analysis provided a discordant result. The results for 158
    children (58 high lead children and 100 low lead children) were
    presented. Thirty-nine non-lead variables were scaled and coded, and
    those which were found to differ between the lead groups at the 10%
    level were controlled as covariates. The covariates controlled were
    father's socioeconomic status, mother's age at subject's birth, number
    of pregnancies, mother's education and Peabody Picture Vocabulary

    score. After this was done the children in the high lead group
    performed significantly less well on the IQ test with a 4´ point
    difference between the groups.

    b)   Winneke et al. (1983)

         A total of 115 children aged 7-12 years living in the vicinity of
    a lead/zinc smelter in Stolberg, Germany, who had provided at least
    one incisor tooth, were investigated. Whole tooth lead measures ranged
    from 2 to 32 µg/g, with a geometric mean of 6 µg/g. There was a
    marginally significant negative association between lead level and
    (short-form) verbal IQ after taking age, gender, perinatal risk
    factors and social confounding factors into account. There was a
    difference of 4.6 IQ points when children with tooth lead levels above
    10 µg/g were compared with those with levels below 4 µg/g. Significant
    negative associations were also observed between tooth lead and the
    error score of Bender Gestalt shape copying test and performance on
    the Vienna reaction time device.

    c)   Smith et al. (1983); Pocock et al. (1987)

         The target population for the study by Smith et al. (1983) (also
    known as the Institute of Child Health/University of Southampton
    study) was 6875 children aged 6-7 years attending schools in one of
    three areas in or near London. In all, 7407 teeth were donated by 4105
    children. Whole tooth analysis of intact incisor teeth donated by
    eligible children was carried out and children whose tooth lead level
    (expressed on an ashed weight basis) was in the lowest percentile
    (< 2.5 µg/g), approximately at the 50th percentile (5-5.5 µg/g) or
    above the 90th percentile (> 8.0 µg/g) were invited to participate.
    The sample was stratified by social grouping. Of the 432 children
    selected, 403 participated in the study. Parental interviews were
    carried out in the home, and information on a wide range of social and
    familial variables was collected. There was a 5 point difference in IQ
    scores between the high and low lead groups before controlling for any
    covariates, but this was reduced to 2 points once covariates
    associated with IQ score were controlled. The authors concluded that
    social factors explain the differences in test performance to such a
    degree that the small differences that remain may be due to other
    social factors not measured. The study underlined the importance of
    the social environment for the child's performance and recognized the
    importance of the relationship between body lead burden and social
    disadvantage.

         Blood samples were taken from 93 children 3-6 months after
    shedding a tooth and after psychometric testing was completed. The
    range of PbB levels in this group was 0.336-1.296 µmol/litre
    (7-27 µg/dl), except for one measurement of 2.064 µmol/litre
    (43 µg/dl), and the mean was 0.614 µmol/litre (12.8 µg/dl). The
    correlation of PbB and PbT was 0.5.

         A reanalysis of this data by Pocock et al. (1989) employed
    multiple regression techniques, using an "optimal" regression
    strategy. The results were comparable to those previously reported.
    Interactional analyses showed no evidence of any differential
    association with lead in different social groups, but revealed a
    significant interaction between PbT and child gender. Significant
    negative association between PbT and IQ was found in boys but not in
    girls.

    d)   Fergusson et al. (1988a,b,c)

         This study was carried out in Christchurch, New Zealand, on a
    sub-sample of a birth cohort of 1265 children born in 1977. Teeth were
    collected from 1035 children, and a dentine "chip" was analysed to
    provide a PbT estimate for 996 children. The mean dentine lead level
    was 6 µg/g. Children from disadvantaged backgrounds had higher mean
    lead levels than those from more advantaged backgrounds, and this
    difference persisted when other social environmental factors (such as
    the effects of housing, proximity to main roads, and childhood pica)
    were taken into account. A WISC-R and reading test were administered
    to children at ages 8 and 9 years, and results were given for 724
    children at age 8 and 644 children at age 9. Small correlations
    between IQ scores and dentine lead were found, but control for
    confounding variables reduced these to statistical non-significance
    for both ages (a correlation of -0.03 to -0.05). Parental IQ was
    assessed. Control for pica as a test of reverse causality further
    reduced the associations between outcomes and PbT.

    e)   Hansen et al. (1989); Lyngbye et al. (1990a,b); Grandjean et al. 
         (1991)

         Children entering schools in the town of Aarhus, Denmark, were
    asked to donate shed teeth. Of a potential sample of 2414 children,
    1291 (54%) provided usable teeth. A sample of circumpulpal dentine was
    obtained from each tooth and analysed by AAS. The mean lead level in
    dentine was 10.7 µg/g, with a log normal distribution and a range from
    0.40 to 168 µg/g. The correlation of whole tooth lead and circumpulpal
    dentine measures were low. Average circumpulpal dentine levels were
    about five times higher than average whole tooth measures. Subsequent
    analysis of PbB from samples obtained two to three years later showed
    an overall mean PbB level of 0.25 µmol/litre (5 µg/dl), with a
    geometric mean of 0.28 µmol/litre (range 0.08-0.63 µmol per litre or
    1.66-13.1 µg/dl) in children selected for the high lead group, and a
    mean of 0.18 µmol/litre (range 0.08-0.70 µmol/litre or mean
    3.74 µg/dl, range 1.66-14.56 µg/dl) in children selected for the low
    lead group, which confirms this as a low exposure group. Children
    whose level of lead in circumpulpar dentine was above 18.7 µg/g (8%,
    N = 110) were selected for the high lead group, and they were matched
    on gender and socioeconomic status (SES) with children whose
    circumpulpar dentine levels were below 5 µg/g. A comprehensive list of
    exclusions, including neurological and medical conditions, atypical

    social situations, and those who did not wish to take part, reduced
    the sample to 156 children for whom results were presented. Initial
    analysis (by matched pair t-test) showed a difference of 6 IQ points
    on full-scale WISC scores, with a difference of nearly 9 points in
    verbal IQ scores. The association of IQ measures with circumpulpar
    dentine lead levels was evident across different socioeconomic groups.
    Confounders associated with circumpulpar dentine lead levels (such as
    pregnancy variables, SES, mother's educational status and child's
    gender) were entered into a stepwise regression model, and explained
    only a small portion of the variance. Lead levels in the dentine
    accounted for a significant (2.8%) of the total variance in full scale
    IQ. Parental IQ was not controlled.

    f)   Rabinowitz et al. (1991)

         From 764 eligible children (attending grades 1 to 3 in 7 selected
    schools in Taiwan), 940 deciduous teeth were collected, the majority
    being incisors. The average age of students was 6.7 years. Two of the
    schools were near lead smelters. Lead levels were determined in 862
    teeth from 692 children using the Boston method which analysed a
    "dentine chip" (Rabinowitz et al., 1989). The mean value was 4.3 µg/g
    of dentine. In all, 493 children were tested using the Ravens Coloured
    Progressive Matrices Test (CPM) and the IQ equivalents were related to
    tooth lead levels. Confounders were selected from a set of 40, based
    on associations with lead and IQ. After correction for seven
    confounding variables (social and perinatal factors), the highly
    significant raw exposure-effect association (r = -0.32) was markedly
    reduced for the full sample to statistical insignificance, although a
    significant lead-CPM association remained for the subsample of girls
    (b = -1.8; SE = 0.78; P < 0.02).

    8.3.4.2  Blood lead studies

    a)   Winneke et al. (1985)

         In a further study by this group carried out in the lead/zinc
    smelter town of Nordenham, Germany, 114 out of 378 children (46%) for
    whom cord PbB levels were available were tested at 6-7 years of age.
    The range of cord PbB levels was 0.19-1.49 µmol/litre (4-31 µg/dl) and
    for current PbB levels was 0.19-1.10 µmol/litre (4-23 µg/dl); the
    means for both perinatal and current PbB levels were 0.39 µmol/litre
    (8.2 µg/dl, SD = 1.6). After correction for confounders by means of a
    stepwise multiple regression analysis, few significant associations
    between PbB levels and performance were observed. There was a marked
    influence of social background factors, but little influence on
    short-form full IQ score (current PbB accounted for 0.3% additional
    variance), and a borderline decrease in performance IQ (PbB accounting
    for 2.4% of the variance). In general, the associations were rather
    stronger with current PbB levels than with cord PbB measures. The
    strongest associations were found with the error scores of the
    difficult version of the Vienna reaction time task.

    b)   Harvey et al. (1984)

         From 483 eligible children from the city of Birmingham, United
    Kingdom, 187 out of 284 contacted were examined at 30 months of age.
    The average PbB level from venous samples was 0.75 µmol/litre
    (15.6 µg/dl, SD = 4). Cognitive ability was assessed by means of four
    tests from the British Ability Scales (BAS) and three tests from the
    Stanford-Binet Intelligence Scale. The raw correlation between PbB
    level and IQ was -0.17 (P < 0.05). Subsequent multiple regression
    analysis with a predetermined set of confounders and a sample size of
    48 yielded an insignificant association between lead and IQ.

    c)   Harvey et al. (1988)

         In a second study, 201 out of 337 eligible children, aged 5.5
    years, from the inner city area of Birmingham, United Kingdom, were
    studied. The mean PbB level was 0.614 µmol/litre (12.8 µg/dl, SD = 4).
    No significant associations between PbB level and IQ (WPPSI) were
    observed after correction for confounding factors. Apart from some
    statistically significant associations between PbB level and reaction
    time, none of the remaining tests showed a significant correlation
    with PbB level after confounder-control.

    d)   Yule et al. (1981)

         Children (N = 166), aged 6 to 12 years, living in the vicinity of
    a leadworks in outer London were examined. PbB levels from venous
    samples ranged from 0.336 to 1.58 µmol/litre (7 to 33 µg/dl, geometric
    mean = 0.648 µmol/litre or 13.5 µg/dl). After controlling for social
    class as the only confounder, significant negative associations were
    found for IQ (WISC-R; full-scale- and verbal-, but not performance-
    IQ). An average IQ difference of 7 points was found when comparing
    children with PbB levels of 0.62 µmol/litre (13 µg/dl) or more with
    those of 0.576 µmol/litre (12 µg/dl) or less. Significant inverse
    associations were also found between PbB level and scores on tests of
    attainment, i.e. reading and spelling but not mathematics.

         A group of 9-year-old children from a middle class area of London
    were studied. Subjects were divided into low PbB (N = 80 and PbB of
    0.336-0.576 µmol/litre, 7-12 µg/dl) and high PbB (N = 82 and PbB of
    0.624-1.15 µmol/litre, 13-24 µg/dl) groups for testing the association
    between PbB level and IQ. No significant association between PbB level
    and IQ was found before or after controlling for covariates (Lansdown
    et al., 1986).

    e)   Schroeder et al. (1985)

         One hundred and four children from a high risk population, aged
    10 months to 6.5 years and with PbB levels ranging from 0.288 to
    2.83 µmol/litre (6-59 µg/dl) were tested for intellectual development
    using age-appropriate tests. After control for social confounding a

    significant negative association was found between PbB level and IQ.
    Fifty children of this cohort were reassessed 5 years later when all
    PbB levels were below 1.44 µmol/litre (30 µg/dl). Neither initial nor
    current PbB level was significantly related to later IQ.

    f)   Hawk et al. (1986)

         In an effort to replicate the above-mentioned findings, 75
    children from the same group, then aged 3 to 7 years, and having a
    mean PbB level of 0.998 µmol/litre (20.8 µg/dl), range
    0.29-2.256 µmol/litre (6-47 µg/dl), were examined using the
    Stanford-Binet test. There were no significant interactions between
    PbB level and factors including age, sex, maternal IQ, quality of the
    care-giving environment (HOME), or socioeconomic status. There was a
    statistically significant negative association between PbB level and
    IQ, although the characteristics of the regression coefficient for
    lead depended on which covariates were included in the model. For the
    final regression model that best controlled for confounding with the
    greatest precision, i.e. the most accurate and precise model,
    containing lead, maternal IQ, HOME and gender, highly significant
    inverse associations were found between both mean and maximum PbB
    levels and IQ.

    g)   Fulton et al. (1987)

         Five hundred and one children, aged 6-9 years, out of 1210
    eligible children from 18 primary schools in central Edinburgh,
    Scotland, were tested for an association between PbB level, IQ and
    attainment. The sample comprised all children in the top quartile of
    the PbB distribution, and a random subsample (approximately 1 in 3) of
    the remainder. The geometric mean PbB level was 0.55 µmol/litre, range
    0.16-1.63 µmol/litre (11.5 µg/dl, range 3.3-34.0 µg/dl). After
    correction for 31 potential confounders a significant negative
    association between PbB level and IQ (BAS) was found. There was a
    dose-response relation with no evidence of a threshold. For purposes
    of statistical analyses, children were placed into 10 groups of about
    50 each on the basis of PbB level. The mean PbB level was
    0.27 µmol/litre (5.6 µg/dl) in the lowest and 1.06 µmol/litre
    (22.1 µg/dl) in the highest group; there was a difference of 5.8 IQ
    points between the two groups. The size of the effects of lead was
    small relative to that of the other factors (0.9% of a total of 45.5%
    explained by the full model).

    h)   Hatzakis et al. (1989)

         Five hundred and thirty-three children out of 1038 eligible from
    four primary schools in the vicinity of the old lead-mining and
    lead-smelting industrial complex of Lavrion in Greece were examined
    neuropsychologically. The mean PbB level from venous samples was
    1.14 µmol/litre, range 0.355-3.067 µmol/litre (23.7 µg/dl, range

    7.4-63.9 µg/dl). In order to control for confounding factors, several
    regression models were evaluated. Starting with a set of 24 potential
    confounders, an optimal model containing 17 potentially confounding
    variables was finally developed for testing the PbB/IQ association.
    The continuous PbB variable was divided into five equally wide classes
    from equal/below 0.715 µmol/litre (14.9 µg/dl, low) to equal/above
    2.16 µmol/litre (45.0 µg/dl). After correction for confounders, a
    highly significant negative association was found between PbB level
    and IQ (WISC-R). A persistent decrease in IQ was only observed at PbB
    concentrations above 1.2 µmol/litre (25 µg/dl). The adjusted
    full-scale IQ difference between "high" and "low" PbB children was 9.1
    points. Highly significant negative associations with PbB were also
    found for the Bender Gestalt test and the Vienna Reaction Device,
    without any indication of a threshold.

    i)   Silva et al. (1988)

         Five hundred and seventy-nine socioeconomically advantaged
    11-year-old children with an average PbB level of 0.53 µmol/litre
    (11 µg/dl), range 0.19-2.4 µmol/litre (4-50 µg/dl) were tested in
    Dunedin, New Zealand. Because the crude inverse association between
    blood lead and IQ was not statistically significant, subsequent
    multivariate analyses were not carried out.

    j)   European multicentre study

         The results from a multicentre study in Europe were reported by
    Winneke et al. (1990). The study linked eight institutions from eight
    European countries; four of the individual study groups were from
    areas near smelters and the others were more general populations. A
    common study protocol with inherent quality assurance elements was
    developed to achieve comparability between the individual studies. In
    all, 1879 children, aged 6-11 years, were studied; however, for some
    of the tests the sample size was reduced to only 971 children. The PbB
    level ranged from 0.24 to 2.88 µmol/litre (5 to 60 µg/dl). Overall
    statistical evaluation was done using a uniform predetermined
    confounder model containing age, gender, occupational status of the
    father, and mother's education. The inverse association between PbB
    level and IQ (four subtests from the WISC-R) was of only borderline
    statistical significance. An IQ decrement of about 3 points was
    calculated for a PbB increase from 0.24 to 0.96 µmol/litre (5 to
    20 µg/dl). The associations between the error scores on the Bender
    Gestalt Test and the Vienna Reaction Device were statistically
    significant and more consistent across study groups, although the
    outcome-variance explained that lead never exceeded 0.8%. The data did
    not allow for the identification of a threshold.

    8.3.4.3  Follow-up studies

    a)   Bellinger et al. (1984)

         A follow-up of some of the children investigated by Needleman et
    al. (1979) was conducted when the children were approximately eleven
    years old. Twenty-two of the original elevated lead group and 48
    children from the low lead group were traced, as well as a group of 52
    children who had not been tested previously, but whose lead levels in
    a "dentine chip" (from teeth analysed at the time of the previous
    study) fell into the mid-range between the high and low lead groups.
    Children in the elevated lead group were significantly younger than
    those in the other two groups. Information on IQ from group-
    administered Otis Lennon Mental Ability tests, conducted 1 to 2 years
    previously, was collected from school records for 15, 52 and 34
    children from the elevated, mid-range and low groups, respectively.
    Scores were inversely related to previously obtained lead measures,
    with a 7-point difference in mean scores between the high and low lead
    groups (t-ratio = 1.65, P = 0.1). No current measures of lead status
    were available.

    b)   Winneke et al. (1989)

         Out of 114 children first tested at age 6 (Winneke et al., 1985)
    76 were retested at age 9. The range of PbB levels had been
    0.187-1.09 µmol/litre, geometric mean = 0.39 (3.9-22.8 µg/dl,
    geometric mean = 8.2) at age six and was 0.21-1.027 µmol/litre,
    geometric mean 0.37 (4.4-21.4 µg/dl, geometric mean 7.8) at age 9
    (r = 0.82). After correction for the same set of confounders used at
    age 6 (age, gender, social background), most of the findings that had
    been found to be significant at age 6, i.e. those related to reaction
    performance (Vienna Reaction Device), remained virtually unchanged
    three years later. The authors concluded that this observation cannot
    necessarily be taken to indicate persistence, because internal
    exposure had essentially remained unchanged over three years.

    c)   Needleman et al. (1990)

         A follow-up assessment was made of 132 young adults first
    investigated by Needleman et al. (1979) at the age of 6-7. Those
    retested were not representative of the group of 270 tested 11 years
    earlier, as they had slightly lower childhood dentine lead levels,
    came from higher SES families, and were more likely to be female.
    Current PbB level was assessed from 48 subjects and none was above
    0.336 µmol/litre (7 µg/dl). A behavioural evaluation including a word
    reading test, and a neurobehavioural assessment was administered to
    each subject, and a self-report of delinquency was obtained. In
    addition school records were reviewed. Multiple regressions indicated
    that higher levels of dentine lead in childhood was associated with
    lower reading scores, lower class rank, increased absenteeism, and
    poorer performance on some of the neurological tests in young

    adulthood. Comparing subjects with dentine lead levels greater than
    20 µg/g with the lowest lead group, children in the high lead group
    were much more likely to drop out of school (unadjusted odds ratio
    4.6; adjusted odds ratio 7.4, CI 1.4-40.8) and to have a reading
    disability (unadjusted odds ratio 3.9; adjusted odds ratio 5.8, CI
    1.7-19.7). Adjustment for covariates made little impact on most of the
    outcomes.

    8.3.4.4  Conclusions and limitations of cross-sectional studies

         A number of general issues are evident from looking at the
    cross-sectional studies as a group. In most studies a negative
    association between lead measures and IQ measures is found in
    uncontrolled data. This difference is usually in the range of 4 to 6
    IQ points and most marked in verbal IQ. Most studies also confirmed
    the positive association between lead measures and indicators of
    social disadvantage, whether this was indicated by SES, maternal
    education, or other more detailed indicators of non-optimal
    child-rearing environments, such as marital quality or maternal
    depression. When these social and other confounding factors are
    controlled, the effect has, in most cases, been to reduce the strength
    of the association between lead measures and IQ, although it remains
    in the same direction. When maternal intelligence has not been
    controlled, the impact on the association between lead and IQ measures
    of correction for covariates tends to be smaller. Where more detailed
    social measures have been controlled, the impact on the magnitude of
    the lead-IQ association has been greater. Associations have, in
    general, been stronger with verbal IQ than with performance IQ.

         The cross-sectional nature of these studies, and also the fact
    that many used a single measure of current exposure, limited their
    usefulness in answering questions relating to the natural history of
    the association between lead exposure and outcomes, including whether
    there were critical periods of exposure, and whether lead associated
    deficits were persistent or reversible.

         Findings from follow-up studies are difficult to interpret.
    Needleman et al. (1990) interpreted their findings as indicating the
    persistence of the effects of lead. Identifying persistence or
    irreversibility in this context has many problems, due to the
    stability of a number of other (non-lead) factors. In addition,
    sampling biases, or failure to control for important covariates in the
    initial or follow-up stages may lead to apparent persistence of an
    effect. A further difficulty is that it is unusual to find a cohort
    where exposure profiles permit easy assessment of persistence.

         Neurobehavioural effects detected at age seven or later usually
    persist in later follow-up studies conducted in other (non-lead) areas
    of research, and for this reason it is more likely than not that
    lead-associated deficits detected during childhood will be detectable
    later. However, there are not enough data to conclude that this is the

    case. Animal data, in which persistence of effects has been shown
    after cessation of exposure, do provide support for the
    irreversibility of lead-induced behavioural toxicity.

    8.3.5  Prospective epidemiological studies on children

    8.3.5.1  Common elements

         The international prospective cohort studies shared a common
    design of pre- or perinatal recruitment, with the principal measure of
    prenatal exposure being the lead concentration in whole blood taken
    antenatally or at birth. To facilitate later comparison of results,
    meetings have been held at which common protocols (within the
    constraints of local conditions) were agreed (Bornschein & Rabinowitz,
    1985; Smith, 1989).

         The protocols were similar in terms of the instruments of
    neurobehavioural assessments, which were well standardized and
    validated for the populations to which they were applied, and although
    the frequency of assessment varied between studies, it was agreed that
    all studies would make an assessment during infancy, in the later
    pre-school period and, if possible, in the school age years.

         All studies conducted several assessments of PbB concentration
    but at varying frequency, and all studies participated in some form of
    quality assurance/quality control (QA/QC) programme for the assessment
    of lead exposure. QA/QC procedures were also used to establish
    inter-examiner reliability (where more than one examiner was used) for
    psychometric and covariate assessments. The importance of such
    procedures was recently discussed (SAHC, 1993).

         In assessing dose-effect relationships between lead exposure and
    outcome variables, most studies used several indices of exposure,
    including both measures of PbB concentration at particular ages and
    average measures of exposure within the lifetimes of the study
    subjects.

         All studies used data analysis procedures which took simultaneous
    account of covariates selected for statistical and substantive
    considerations.

    8.3.5.2  Study descriptions

    a)   Boston study

         The study population was recruited from 11 837 infants born at
    the Brigham and Women's Lying-In hospital in Boston between April 1979
    and April 1981 (Bellinger et al., 1984). On the basis of cord PbB
    levels of approximately 2500 children, the 90th, 50th and 10th

    percentiles were identified and used as eligibility criteria for
    enroling 249 children in three exposure groups: low (< 0.14 µmol/litre
    or 3 µg/dl, N = 85), medium (0.288-0.336 µmol/litre or 6-7 µg/dl,
    N = 88) and high (> 0.48 µmol/litre or 10 µg/dl, N = 76). No child
    had a cord PbB level greater than 1.2 µmol/litre (25 µg/dl).

         Exclusion criteria included birth complications or medical
    conditions associated with developmental difficulties (e.g.,
    gestational age < 34 weeks, Down's syndrome), a non-English speaking
    family or residence in an area considered unsafe for home visitors.
    The cohort consisted largely of children from intact middle and
    upper-middle class families. Approximately 85% of the children were
    white.

         Capillary blood collected at ages 6, 12, 18 and 24 months was
    analysed for lead by ASV. Venous blood collected at ages 57 months and
    at 10 years was analysed for lead by AAS. In contrast to other
    prospective cohorts, the mean PbB level did not rise postnatally and
    indeed never exceeded 0.384 µmol/litre (8 µg/dl).

         The principal psychometric tests included the Bayley Scale of
    Infant Development at 6, 12, 18 and 24 months, the McCarthy Scales of
    Children's Abilities at 57 months and the Wechsler Intelligence Scale
    for Children - Revised and the Kaufman Test of Educational Achievement
    at age 10 years. After age 2 years, the mean level of performance was
    above the expected population mean (e.g., mean full scale IQ at age 10
    years = 119) (Bellinger et al., 1992).

         A variety of potential confounding factors was assessed including
    maternal IQ, the quality of rearing environment (HOME scores at 6, 24,
    57 and 120 months) and prenatal, perinatal and postnatal medical and
    socio-demographic factors. 

    b)   Cincinnati study

         Women attending prenatal clinics in predesignated leadhazardous
    residential areas of Cincinnati, Ohio, USA, were consecutively
    recruited from 1979 to 1984 (Dietrich et al., 1987) Prenatal
    exclusions were for maternal prenatal alcohol and drug abuse,
    psychosis, diabetes, mental retardation; neonatal exclusions included
    birth weight under 1.5 kg and gestational age (by physical
    examination) below 35 weeks. Furthermore, recruited infants had to
    have an Apgar score of 6 or more at 5 min postpartum and to have no
    serious medical conditions. Of the mothers, 87% were single and 86%
    were receiving public assistance. The final follow-up sample of 305
    infants who attended their second clinic visit at 3 months of age was
    85% African-American and 50% female.

         Mean prenatal maternal PbB levels as assessed by ASV were
    0.38 ± 0.177 µmol/litre (8.0 ± 3.7 µg/dl). Mean neonatal PbB levels
    were also low at 0.221 ± 0.134 µmol/litre (4.6 ± 2.8 µg/dl). Following

    the development of prewalking progression and normal hand-to-mouth
    behaviours, PbB levels began to rise, peaking at around 21 months with
    a mean of approximately 0.89 µmol/litre (18 ± 6 µg/dl). Approximately
    35% of the sample had at least one PbB concentration equal to or
    greater than 1.2 µmol/litre (25 µg/dl) sometime during the first 5
    years of life, while 79% exceeded 0.72 µmol/litre (15 µg/dl) during
    the same period (Dietrich et al., 1991, 1992). Virtually all children
    (95%) exceeded 0.48 µmol/litre (10 µg/dl) during the first 5 years of
    life. Postnatal PbB concentrations were assessed on a quarterly basis
    beginning at 10 days postpartum. The vast majority of blood samples
    were collected by venepuncture.

         Major neurobehavioural assessments included the Bayley Scales of
    Infant Development administered at 3, 6, 12 and 24 months of age. The
    Kaufman Assessment Battery for Children was administered at 4 and 5
    years. A comprehensive examination of gross and fine neuromotor
    functions (the Bruininks-Oseretsky Scales) was administered at 6
    years. Finally, performance on the Wechsler Intelligence Scale for
    Children - Revised was assessed following school entry at 6.5 years.

         Intellectually, this was a low functioning cohort with a mean
    full-scale IQ at 6.5 years of 86.9 ± 11.3 points (Dietrich et al.,
    1993a).

         Covariates measured included assessments of obstetrical and
    perinatal complications, birth anthropometrics, tobacco and alcohol
    consumption, observational assessments of the quality of care-taking
    in the home at 6, 12, 24 and 36 months, social class, iron status and
    various aspects of child health likely to affect neurobehavioural
    performance (e.g., otitis media, sensory deficits, allergies, seizure
    disorders). Maternal IQ was assessed postnatally. All medical and
    psychometric testing was conducted at a single welfare clinic located
    in the heart of the recruitment area.

    c)   Cleveland

         Five hundred and forty-three infants delivered to women at a
    large inner city general hospital between February 1981 and March 1982
    were considered for inclusion in this study of the joint or
    independent effects of fetal alcohol exposure and lead on child
    development (Ernhart et al., 1985). Due to pre-term birth or illness,
    16% were excluded from further study. Also excluded were mothers
    reporting the use of narcotics and those with identifiable
    psychological disorders. The sample was predominantly white (65%) and
    of lower SES. Approximately 50% of mothers were alcoholics.

         PbB concentrations were assessed by AAS (208 mothers at delivery
    and in 178 cords resulting in 142 mother-infant pairs with complete
    prenatal lead exposure data). Postnatally, blood samples were analysed
    for approximately half the birth cohort at ages 6 months, two years
    and three years of age. At least one perinatal or postnatal PbB

    concentration was available for each of 285 children with
    developmental data past the neonatal period (Ernhart et al., 1987,
    1989a,b).

         Mean maternal PbB concentrations (N = 185) were low
    (0.31 ± 0.086 µmol/litre or 6.5 ± 1.8 µg/dl) and mean cord PbB
    concentrations (N = 162) were also low (0.29 ± 0.10 µmol/litre or
    6.0 ± 2.1 µg/dl). Postnatally, mean PbB concentrations (N = 151) were
    0.48 ± 0.16 µmol/litre (10.0 ± 3.3 µg/dl) at 6 months,
    0.80 ± 0.28 µmol/litre (16.7 ± 5.9 µg/dl) at 2 years (N = 165) and
    0.80 ± 5.9 µmol/litre (16.7 ± 5.9 µg/dl) at 3 years. The lowest and
    highest PbB levels recorded postnatally were 0.24 and 2.016 µmol/litre
    (5 and 42 µg/dl), respectively.

         Neurobehavioural assessment included selected sub-scales of the
    Brazelton Neonatal Behavioural Assessment Scales and Graham Rosenblith
    Behavioural Examination of the Neonate at > 24 h postpartum, the Kent
    Infant Development scale at 6 months, the Bayley Scales of Infant
    Development at 6, 12 and 24 months, the Stanford-Binet Intelligence
    Scale at 3 years, and finally the Wechsler Preschool and Primary Scale
    of Intelligence at approximately 5 years of age. All assessments
    beyond the neonatal period were conducted in subjects' homes.

         Intellectual attainment in this high risk, socioeconomically
    disadvantaged cohort was typically low as shown by a mean full-scale
    IQ at approximately 5 years of 87.5 ± 16.6 points (Ernhart et al.,
    1989a).

         In addition to the usual assessments of obstetrical/perinatal
    complications and neonatal status, other covariates measured included
    the Michigan Alcoholism Screening Test, the Peabody Picture Vocabulary
    Test as an assessment of maternal intelligence, and the Authoritarian
    Family Ideology Scale. Women were also questioned as to their
    consumption of alcohol and tobacco during pregnancy. Observational
    assessments of care-taking quality in the home were conducted at 1, 2,
    3 and 4 years.

    d)   Glasgow study

         The study group for CNS outcomes consisted of 151 subjects drawn
    from an initial sample of 885 families exposed to various levels of
    dietary lead, principally due to a plumbosolvent water supply (Moore
    et al., 1989). All subjects were born in the United Kingdom and spoke
    English as a first language. The sample was divided into three groups
    of approximately equal numbers (matched for social class) based on
    maternal prenatal PbB levels: high (> 1.44 µmol/litre or 30 µg/dl),
    medium (0.72-1.2 µmol/litre or 15-25 µg/dl) and low
    (< 0.48 µmol/litre or 10 µg/dl). The SES of the sample ranged from
    the chronically unemployed to the professional classes.

         PbB levels were assessed postnatally at 1 and 2 years of age and
    found to be 0.734 and 0.777 µmol/litre (15.3 and 16.2 µg/dl),
    respectively.

         Measurements of neurobehavioural outcome where the Bayley Scales
    of Infant Development were used were administered at 1 and 2 years of
    age.

         Covariates considered in the data analysis included a measure of
    obstetrical complications, birth weight, birth order, SES of the
    father in postnatal follow-up years 1 and 2, and observational
    assessments of care-taking quality in the first and second years of
    follow-up.

    e)   Kosovo Study

         Five groups of infants from two communities in Kosovo,
    Yugoslavia, were studied (Graziano et al., 1990; Wasserman et al.,
    1992). Mitrovica was the site of a lead smelter, refinery and battery
    plant, and Pristina an area of minimal lead exposure. Three groups
    were recruited from Mitrovica on the basis of cord PbB concentration:
    (1) < 0.72 µmol/litre (< 15 µg/dl, N = 78); (2) 0.72-0.96 µmol/litre
    (15-20 µg/dl, N = 99); (3) > 0.96 µmol per litre (> 20 µg/dl,
    N = 217). Two groups were recruited from Pristina: (4) a group with
    cord PbB levels < 0.72 µmol/litre (< 15 µg/dl) and (5) a group
    matched to children in group (3) in terms of the distribution of
    maternal and paternal education. Grouping was not considered in the
    statistical analysis. The total cohort (N = 541) was largely Albanian
    and Serbian. Infants with CNS defects, chromosomal anomalies, distant
    residence and twin pairs were excluded.

         PbB level and various indices of iron status (EP, Hb, serum
    ferritin) were measured in venous blood collected in midpregnancy,
    delivery and at postnatal ages 6, 12, 18 and 24 months. Mean PbB level
    amongst children in Mitrovica exceeded 0.96 µmol/litre (20 µg/dl) at
    all ages, reaching 1.68 µmol/litre (35 µg/dl) at 24 months. Amongst
    children in Pristina, the mean PbB level was below 0.48 µmol/litre
    (10 µg/dl) at all ages.

         The Mental Scale of the Bayley Scales of Infant Development was
    administered at ages 6, 12, 18 and 24 months, and the Motor Scale at 6
    and 12 months. At 24 months, the mean Mental Development Index in the
    cohort (n = 392) was 105.2 (SD = 18.1).

         Anthropometric measurements and assessments of child and family
    health, diet, and demographic status were carried out at each age. An
    adaptation of the HOME Scale was administered at ages 3 and 4 years
    and maternal IQ was assessed using the Raven's Standard Progressive
    Matrices.

    f)   Port Pirie study

         Port Pirie (population 16 000 in 1979) is an industrial town 200
    km north-west of Adelaide, South Australia, with a large and
    long-standing smelting facility (Wigg et al., 1988). Over the period
    1979-1982, a cohort of 723 children from Port Pirie and surrounding
    rural areas was recruited into the study. The cohort represented a 90%
    sample of all children born in the area during this period.

         Lifetime lead exposures were estimated from venous PbB
    concentrations obtained antenatally (at 14-20 weeks of gestation, and
    32 weeks), at delivery (maternal and cord blood), and from capillary
    samples taken postnatally at ages 6, 15, and 24 months, and annually
    thereafter. Lead concentrations in all samples were determined by AAS.

         At the time that each blood sample was collected, the nurse
    interviewer also conducted structured interviews to obtain information
    on a range of demographic, familial, behavioural, medical and social
    environmental factors. HOME scores were assessed when the child was 3
    and 5 years of age. Maternal IQ was also assessed. The Bayley Scales
    of Infant Development was administered at 2 years, the McCarthy Scales
    of Children's Abilities at 4 years and the Wechsler Intelligence Scale
    for Children - Revised version (WISC-R) at 7 years.

    g)   Sydney study

         A total of 318 children was recruited from among infants born at
    three maternal hospitals between April 1982 and March 1983 (Cooney et
    al., 1989a). Exclusion criteria included prematurity, low birth
    weight, severe medical problem, single or non-English speaking mother
    or maternal drug or alcohol problems. All children were white and
    generally represent a middle-class population.

         PbB concentration was measured in maternal venous blood during
    hospitalisation (geometric mean = 0.388 µmol/litre or 8.1 µg/dl,
    geometric SD = 0.067 µmol/litre or 1.4 µg/dl). Additional blood
    samples were collected at 6-month intervals to age 4 years and at ages
    5 and 7 years. Up to 24 months of age, approximately 50% of the blood
    samples were capillary, while almost all samples collected
    subsequently were venous. Geometric mean PbB levels between 6 months
    and 4 years were in the range of 0.48-0.72 µmol/litre (10-15 µg/dl),
    falling to 0.39 and 0.37 µmol/litre (8.3 and 7.7 µg/dl) at ages 5 and
    7, respectively.

         The Bayley Scales of Infant Development were administered at ages
    6, 12 and 24 months, the McCarthy Scales of Children's Abilities at
    ages 3 (n = 215) and 5 (n = 200), and the Wechsler Intelligence Scale
    for Children - Revised at 7 years (n = 175). All assessments, at least
    until age 5, were conducted in the child's home. The mean Mental

    Development Index and General Cognitive Index (GCI) scores were in the
    intellectual normal range of 107 to 117 with a mean general cognitive
    index at 4 years of 107.3 ± 14.2.

         Covariates measured included maternal IQ, quality of care-taking
    environment, obstetrical and postnatal factors.

    8.3.5.3  Summary of differences between studies

         The prospective studies differed to such an extent that it would
    be surprising if they yielded identical findings. This does not mean
    that the studies cannot be compared, merely that these differences and
    their potential impact on study findings must be acknowledged in any
    assessment of the studies as a group (Bellinger & Stiles, 1993). There
    are substantial differences in the degree of confounding between lead
    exposure and the other correlates of poor developmental outcomes. For
    example, mean maternal IQ ranges from 121 (Boston) to 74-75
    (Cleveland, Cincinnati), reflecting vast cohort differences in
    socioeconomic standing and child-rearing context.

         The studies also varied substantively in terms of the lead
    exposure profiles observed. Site differences resulted in considerable
    variability in postnatal lead exposures depending on exposure patterns
    (residence near primary lead smelters, residence in deteriorated
    housing, proximity to traffic, and quality of domestic
    drinking-water).

         Despite broad similarities in the approaches to data analysis,
    there were substantial differences in the form in which the findings
    were reported, impeding direct comparison of study results. Results
    were diversely reported in terms of partial correlation coefficients
    and standardized and unstandardized regression coefficients, and there
    was variable use of raw (untransformed) and log exposure measures.
    Most, but not all studies presented their results in a manner
    conducive to some kind of quantification of dose-effects
    relationships.

         Clearly the variation in numbers of participants available to
    individual studies (both initially, and as a consequence of attrition)
    has a direct effect on the statistical power of each study - which is
    also affected by the degree of confounding.

    8.3.5.4  Results of studies

    a)   Infancy and early preschool assessments

         The pattern of findings in infant and early preschool assessments
    for which the principal instrument of assessment was the Bayley Scales
    of Infant Development is inconsistent. Using different indices of
     in utero exposure (e.g., prenatal, at delivery, cord or neonatal PbB
    levels), three of the prospective studies found associations with

    slower sensorimotor development up to 6 months or 1 year of age
    (Ernhart et al., 1987; Dietrich et al., 1987, 1989) or up to 2 years
    of age (Bellinger et al., 1987), but these tended to attenuate over
    time (Dietrich et al., 1990; Bellinger et al., 1991a). Other studies
    did not observe statistically significant adjusted relationships of
    prenatal exposure with slower sensorimotor development (Moore et al.,
    1989; Wigg et al., 1988; Cooney et al., 1989a,b; Wasserman et al.,
    1992).

         Two studies observed an inverse association between postnatal
    lead exposure on mental development in later infancy (Wigg et al.,
    1988; Wasserman et al., 1992), although in one study it was PbB level
    at 6 months of age (Wigg et al., 1988) and in the other study it was
    the level at 24 months (Wasserman et al., 1992) that was most
    predictive.

    b)   Later preschool and school age assessment

         Despite some inconsistency in the pattern of findings in infants,
    there has been a convergence of positive findings on later
    neurobehavioural outcomes in the prospective studies. This may
    reflect:

    1)   the greater reliability and precision of measurement attained
         with assessments of the older child; or

    2)   an effect of lead on abilities that cannot easily be tested
         during infancy (e.g., executive, regulative and organizational
         skills, higher order reasoning).

    c)   Boston study

         Among 169 children assessed at age 57 months, the major finding
    was an association between PbB level at 2 years of age and the GCI of
    the McCarthy Scales. There was a decrease of 2.95 ± 1.42 GCI units for
    an increase in PbB of 0.48 µmol/litre (10 µg/dl). Since the group had
    an average PbB level of 0.336 µmol/litre (7.0 µg/dl) over 2 years,
    this decrement was relevant to a range of approximately
    0.19-0.67 µmol/litre (4-14 µg/dl) Bellinger et al., 1991a).

         At 10 years of age, 148 children were reassessed. PbB
    concentration at 2 years of age was inversely associated with the
    WISC-R Full Scale IQ. Each 0.48 µmol/litre (10 µg/dl) increase was
    associated with a 5.8 point decline in IQ (95% CI 1.7 to 9.9)
    (Bellinger et al., 1992).

         PbB concentration at 24 months was also inversely related to the
    Battery composite score in the Kaufman Test of Educational Achievement
    (Brief form) (-0.89 ± 0.24). The skills assessed were mathematics,
    reading and spelling.

         The point estimates associated with other postnatal ages were
    generally in the same direction but only PbB level at 2 years revealed
    conventional statistical significance. Other specific
    neuropsychological tests did not reveal a clear pattern of
    neuropsychological deficit (Stiles & Bellinger, 1993).

    d)   Cincinnati study

         The Kaufman Assessment Battery for Children was administered to
    approximately 260 children at 4 and 5 years of age (Dietrich et al.,
    1991, 1992). The principal findings at 4 years were that higher
    neonatal PbB levels were associated with poorer performance on all
    Kaufman subscales. However, this inverse association was limited to
    children from the poorer families. Following full covariate
    adjustment, few statistically significant associations between
    postnatal PbB levels and Kaufman scales could be found. However, the
    results did suggest a weak relationship between postnatal PbB levels
    and performance on a Kaufman subscale which assesses visual spatial
    and visual-motor integration (adjusted regression coefficient -0.12 SE
    not published).

         At 5 years of age, postnatal PbB levels were associated with
    poorer performance an all subscales of the Kaufman Battery. However,
    after adjustment for covariates, few statistically significant
    relations remained. Nevertheless, as at 4 years of age, the subscale
    assessing visual-spatial and visual-motor skills was most sensitive,
    with average lead exposure during the 4th year of life being
    significantly associated with performance (-0.12 units per
    0.048 µmol/litre or per µg/dl, SE not published).

         At the age of 6.5 years, 253 children in the Cincinnati cohort
    were administered the WISC-R (Dietrich et al., 1993a). The major
    findings were that postnatal PbB concentrations were inversely
    associated with full-scale IQ and performance IQ. Following
    statistical adjustment for covariates, including maternal IQ and
    quality of home care, a statistically significant relationship was
    retained between postnatal PbB concentrations at nearly all ages (and
    including lifetime averages) and Performance IQ. Further analysis
    revealed that average lifetime PbB concentrations in excess of
    0.96 µmol/litre (20 µg/dl) were associated with deficits in IQ of the
    order of 7 points, when compared with children with mean
    concentrations of less than 0.48 mol/litre (10 µg/dl). The regression
    coefficient for Performance IQ on mean lifetime PbB level was -0.26
    (± 0.12) IQ units per 0.048 µmol/litre or per µg/dl. This corresponded
    to a 2.6 unit decrement associated with a 0.48 µmol/litre (10 µg/dl)
    movement in lifetime average PbB level, for which the average
    (interpolated graphically) is close to 0.67 µmol/litre (14 µg/dl).

         At 6 years of age, the Bruininks-Oseretsky Test of Motor
    Proficiency (BOTMP) was administered to 245 children in the Cincinnati
    Cohort (Dietrich et al., 1993a). Following statistical adjustment for

    covariates, neonatal PbB levels were associated with poorer
    performance on a measure of upper-limb speed and dexterity and the
    fine-motor composite. Postnatal PbB levels also remained significantly
    associated with poorer scores on measures of bilateral coordination,
    visual-motor control, upper-limb speed and dexterity, and the fine
    motor composite.

         Children having an average mean lifetime PbB level of
    approximately equal to or exceeding 0.43 µmol/litre (9 µg/dl) appeared
    to experience a deficit in both fine and gross motor skills relative
    to children in the lowest PbB quartile. Children in the highest
    average lifetime PbB quartile had scores on the gross-motor subtest
    assessing bilateral coordination of approximately 0.5 standard
    deviations (2.5 points) lower than their counterparts in the lowest
    quartile. Children in the highest average lifetime quartile also
    scored more poorly in the fine-motor functioning, having scores of
    approximately 0.6 standard deviations lower (6.3 points) than those in
    the lowest quartile.

    e)   Cleveland study

         The WPPSI was administered to 242 children at the age of 4 years
    and 10 months (Ernhart et al., 1989a). Statistically significant
    correlations between IQ (full scale and subscales) and PbB level
    measured pre- and perinatally and at ages 2 and 3 years (and a
    lifetime postnatal average) became non-significant after adjustment
    for covariates. No estimates for effect size were presented in the
    original report (Ernhart et al., 1989a).

    f)   Glasgow study

         No neurobehavioural assessments were carried out beyond 2 years
    of age (Moore et al., 1989).

    g)   Kosovo study

         No results beyond 2 years of age were reported (Graziano et al.,
    1990; Wasserman et al., 1992).

    h)   Port Pirie cohort study

         The principal finding at 4 years was an inverse association of
    McCarthy Scales of Children Abilities (MSCA) scores (General Cognitive
    Index (GCI), Perceptual Performance and Memory) with most indices of
    lead exposure (McMichael et al., 1988). After adjustment for
    covariates, the association of GCI was not signifi-cant at certain
    specific ages. However, it remained significant for the integrated
    postnatal average for which an effect size of 7.2 GCI points lost in
    association with an increase of lifetime average PbB level from 0.5 to
    1.5 µmol/litre (10.4-31.2 µg/dl) was estimated.

         Between 7 and 8 years of age the IQ of 494 children was assessed
    with the WISC-R (Baghurst et al., 1992). After adjustment for
    covariates there was little association with pre- and perinatal lead
    exposure assessments but significant decrements in full-scale IQ of
    between 3.7 and 4.8 points (depending on age) for each (natural) log
    unit increase in lifetime average PbB concentration were observed. The
    estimated effect size for the lifetime average exposure up to 3 years
    of age was a loss of 5.3 IQ points in association with an increase in
    PbB level from 0.48 to 1.44 µmol/litre (10 to 30 µg/dl).

         The Block Design subscale of the WISC-R, which tests spatial
    abilities, exhibited the strongest association with lead exposure and
    estimated effect sizes were stronger in girls than boys for both the
    MSCA and WISC-R.

    i)   Sydney study

         At the age of 4 years, 207 children were assessed with the
    McCarthy Scales of Children Abilities (MSCA) and virtually no
    significant associations with any measures of PbB were observed
    (Cooney et al., 1989b).

         Follow-up assessments in 175 children using WISC-R at age 7 also
    yielded no associations (crude or adjusted) with blood lead history
    (Cooney et al., 1991).

    8.3.5.5   Questions prospective studies have not answered

         A disappointment of the prospective studies was their inability
    to reach any obvious consensus on the behavioural phenotype associated
    with low-level lead exposure, or on age(s) of critical sensitivity.
    This latter fact may reflect the phenomenon of intra-individual
    "tracking", whereby an individual maintained approximately consistent
    ranking with respect to his or her PbB concentration at any age, or it
    may reflect the need for chronic exposure over extended periods in
    order for causal effects to become apparent.

    8.3.5.6   Attempting a consensus

         As discussed previously, no clear delineation of an age (or age
    range) of maximal sensitivity to lead exposure has emerged from the
    prospective studies, and there is a diversity of reporting styles
    which makes attempts to compare studies and establish a consensus of
    opinion very difficult.

         Nevertheless, it would appear essential to attempt some sort of
    synthesis of studies performed so far, given that the social and
    economic consequences of preventing an adverse health effect are
    potentially enormous.

         A meta-analysis by Needleman & Gatsonis (1990) concluded that the
    overall pattern of results was most unlikely to have occurred by
    chance. As an approach to the problem of obtaining the "best" estimate
    of the effect of lead exposure on neuro-behavioural development, a
    common estimate of the partial correlation between lead burden and IQ
    was determined. However, in order to derive such an estimate with some
    degree of certainty, it is essential to find a common outcome, and to
    identify studies for which at least broadly comparable exposure
    measures are available.

         Schwartz (1993) has identified three prospective studies and five
    cross-sectional studies that examined the association of lead exposure
    as determined by PbB concentrations with full scale IQ. For the
    prospective studies, the exposure measures used were 24-months PbB
    level (Boston) and lifetime average up to 3 years of age (Cincinnati
    and Port Pirie).

         There was only one measure for each of the cross-sectional
    studies. Despite the compromises necessary for such a comparison, the
    estimated effect sizes for an increase in PbB level from 0.48 to
    0.96 µmol/litre (10 to 20 µg/dl) for all seven studies lay between 1
    and 6 IQ points lost (with five of the studies lying between 1 and 4),
    and the associated 95% confidence limits only just embraced the value
    zero for 3 of the 7 studies.

         Thus, a broad consensus does emerge from a rough comparison. From
    the estimates summarized by Schwartz (1993), a weighted mean for the
    six coefficients accompanied by a standard error of 0.4 was estimated
    to be 2.6 IQ units lost for an increase in PbB from 0.48 to
    0.96 µmol/litre (10 to 20 µg/dl).

    8.3.6  Task group overview and interpretation of prospective studies
           on children

    8.3.6.1  Rationale

         As a complement to the narrative review, prepared by the Task
    Group, a quantitative overview of the findings of the prospective
    studies was considered to be a valuable means of assessing the overall
    strength of evidence for an association between blood lead summary
    measures and school age IQ.

         While individual studies vary somewhat in their design and their
    analysis strategies, their underlying objectives are closely related.
    For any generally applicable conclusions on general population lead
    exposure and its effect on intellectual attainment to be reached, some
    consistency of evidence across studies is highly desirable.

         Since any single study has considerable random error in
    estimating a relationship, a quantitative overview (or meta-analysis)
    can be a valuable means of defining what is the plausible magnitude of
    statistical association between PbB measures and child IQ.

         It seems appropriate, given the quite different types of study
    design, to present two separate meta-analyses dealing with prospective
    studies and cross-sectional studies, respectively.

         Any such meta-analysis needs cautious interpretation. The caveats
    in drawing causal inferences from statistical associations apply to
    any meta-analysis just as they do to individual studies. Any increase
    in statistical precision regarding the magnitude of association does
    not eliminate the potential for bias in any observational study on
    human populations.

    8.3.6.2  The prospective studies

         There are four prospective studies: Boston (Bellinger et al.,
    1992); Sydney (Cooney et al., 1989a,b); Cincinnati (Dietrich et al.,
    1993a); and Port Pirie (Baghurst et al., 1992) which presented
    quantitative results in a form suitable for an overview (i.e. with
    regression coefficients and standard errors). The Task Group
    concentrated on school age assessment because:

    a)   the studies used the same outcome measure (WISC-R), which is
         widely accepted and has been previously evaluated in cross-
         sectional studies; and

    b)   intellectual assessment at primary school age (6-10 years) is,
         perhaps, of greatest overall relevance in reaching a consensus on
         the public health importance of childhood lead exposure.

         Unfortunately, the Cleveland study did not publish findings in
    the same quantifiable manner. The potential impact of this apparently
    "negative" study on the overall evidence will be discussed in section
    8.3.6.2(c).

    a)   Measures of exposure, outcome and association

         All four studies have used blood lead as the measure of body lead
    burden, but the timing and frequency of examinations has varied as has
    the choice of summary measures (averages over time).

         In a meta-analysis it seems appropriate to concentrate on
    measures of association after adjustment for confounders. While there
    is some variation in choice of confounders, the two key covariates
    (mother's IQ and HOME scores) appear in all four studies.

         The Port Pirie study used the logarithm of the PbB concentration
    in the multiple regression analysis while the other three studies used
    untransformed PbB concentration. This difference in statistical style
    can be reconciled by converting the lead regression coefficients to
    estimated changes in IQ for a specific increase in PbB concentration,
    i.e. from 0.48 to 0.96 µmol/litre (10 to 20 µg/dl). This equals 10
    times the regression coefficient (as published) for the Boston and
    Cincinnati studies and log 2 times the regression coefficient for the
    Port Pirie study.

         It is recognized that the specific subscales of performance and
    verbal IQ are of separate interest (e.g., the performance IQ has been
    rather consistently shown to be more reliably associated with
    postnatal lead exposure amongst the prospective studies (McMichael et
    al., 1988; Bellinger et al., 1991a; Dietrich et al., 1991, 1992,
    1993b). However, it appears more important to focus on full-scale IQ
    as the primary outcome measure, otherwise concerns about post-hoc
    selection may be raised. Such full-scale IQ was measured at age 6.5, 7
    and 10 years in the Cincinnati, Port Pirie, Sydney and Boston studies,
    respectively.

    b)   Display of individual study findings

         Since most studies have not shown perinatal PbB level to be
    predictive of intellectual performance, attention will be focused on
    the various measures of postnatal PbB level. In fact, there are seven
    such measures in each study that have been related to full-scale IQ in
    the reports. Tables 21 and 22 summarize the magnitude of association
    together with the standard error for all reported analyses, each after
    adjustment for confounders.

         Except for the earliest postnatal PbB values in the Boston and
    Cincinnati studies, the adjusted associations between PbB level and IQ
    are consistently negative (i.e. inverted). However, the "random noise"
    in each study's analysis means that there is considerable variation in
    levels of statistical significance. Hence, there is a need for a more
    systematic overview combining the evidence from all studies.

    c)   A quantitative overview

         The prime difficulty here is the different choices of PbB
    summaries in the four studies. In undertaking a meta-analysis it would
    have been better if all studies had essentially the same PbB measures
    in their multiple regressions. Possible choices might have been (i) a
    specific time point, e.g., 2 years or (ii) the means over a specific
    interval, e.g., 0 to 5 years.

         However, from the available analyses, some broadly similar
    summaries can be chosen, as follows:

        Table 21.  Magnitude of association between blood lead and full-scale IQ
               from four prospective studies
                                                                                  
                   No. of         Age at which blood lead       Estimated change in
                   children       was measured                  full-scale IQ a
                                                                                  

    Boston           148          year 1                            0.0 (1.6)
                                  year 1´                          -1.2 (1.8)
                                  year 2                           -5.8 (2.1)
                                  year 5                           -2.6 (2.9)
                                  year 10                          -4.6 (5.2)
                                  mean of years 2-5                -8.2 (2.8)
                                  mean of years 2-10               -8.6 (3.4)
                                  mean of 6 months to
                                   10 years (unpublished)          -5.7 (3.2)

    Cincinnati       251          mean of year 1                    0.1 (1.4)
                                  mean of year 2                   -0.2 (0.8)
                                  mean of year 3                   -1.3 (0.9)
                                  mean of year 4                   -1.5 (1.0)
                                  mean of year 5                   -2.3 (1.1)
                                  mean of year 6                   -3.3 (1.4)
                                  mean of years 1-6                -1.3 (1.1)

    Port Pirie       490          mean of first 15 months          -2.8 (1.4)
                                  mean of years 1-2                -3.2 (1.5)
                                  mean of years 1-3                -3.3 (1.6)
                                  mean of years 1-4                -3.2 (1.7)
                                  mean of years 1-5                -3.0 (1.7)
                                  mean of years 1-6                -2.8 (1.7)
                                  mean of years 1-7                -2.6 (1.7)

    Sydney           175          mean of years 1 and 2            -2.7 (2.8)
                                  mean of years 3, 4 and 5         -1.9 (1.9)
                                  year 7                           -1.4 (1.7)
                                  mean of years 1-7                -1.6 (2.2)
                                                                                  

    a  From 0.48 to 0.96 µmol/litre (10 to 20 µg/dl) blood lead
       (and its standard error) after adjustment for confounders
    
    Table 22.  Association between blood lead levels and full-scale
               IQ from ten cross-sectional studies
                                                                   
                                  No. of         Estimated change in
                                  subjects       full-scale IQa
                                                                   

    Lavrion, Greece                 509             -2.7 (0.7)

    Edinburgh, Scotland             501             -2.6 (0.9)

    Greenwich, England              129             -5.6 (3.2)

    European, multi-centre

    Bucharest, Romania              301             -0.4 (2.6)

    Budapest, Hungary               254             +0.8 (1.8)

    Modena, Italy                   216             +0.9 (4.2)

    Sofia, Bulgaria                 142             +2.2 (2.3)

    Dusseldorf 1, Germany           109             -4.6 (4.5)

    Dusseldorf 2, Germany           109             -3.9 (5.1)

    Zagreb, Croatia                  48             -1.5 (4.5)
                                                                   

    a  From 0.48 to 0.96 µmol/litre (10 to 20 µg/dl) blood lead
       (and its standard error) after adjustment for confounders


    i) There is some advantage in considering the mean PbB level over a
    number of years, since it summarizes cumulative exposure and also
    achieves a more reliable ranking of individuals than a single value.
    The available long-term means are:

         Boston:        mean of 6 months to 10 years (D. Bellinger,
                        personal communication to the IPCS, 1993)
         Cincinnati:    mean of years 1 to 6
         Port Pirie:    mean of years 1 to 7
         Sydney:        mean of years 1 to 7

         Although they differ in frequency and age range, these three
    summaries should be very highly correlated with one another. The
    resultant meta-analysis is displayed in Fig. 15. Each of the
    individual study confidence intervals includes zero, indicating that
    significance at P < 0.05 was not reached. However, the combined
    evidence, weighting studies according to the inverse of their
    variance, produces a weighted mean decrease in full-scale IQ of 2
    points for a 0.48-µmol/litre (10-µg/dl) increase in PbB level, with
    95% confidence interval from -0.3 points to -3.6 points (P = 0.01).

    ii) An alternative approach is to consider PbB level at a specific
    time or average over shorter intervals of time. For this approach, the
    analyses of these studies are most comparable during the first three
    years, as follows:

         Boston:        2 year measure
         Cincinnati:    mean of year 3
         Port Pirie:    mean of years 1 to 3
         Sydney:        mean of years 1 and 2

         This second meta-analysis is displayed in Fig. 16. The data here
    support an inverse association more strongly than in Fig. 15, and the
    combined evidence estimates a mean decrease of 2.6 IQ points for a
    0.48-µmol/litre (10-µg/dl) increase in PbB level, with 95% confidence
    interval from -1.2 points to -4.0 points (P < 0.001).

         However, it should be noted that the estimates here are heavily
    dependent on the choice of time points for PbB level. For instance, if
    one instead chose:

         Boston:        1.5 year measure
         Cincinnati:    mean of year 2
         Port Pirie:    mean of years 1 and 2
         Sydney:        mean of years 1 and 2

    then the combined estimate is roughly halved in magnitude and is of
    borderline significance.

         One needs to recall that the other prospective study, carried out
    in Cleveland (Ernhart et al., 1989a), has no equivalent data, but
    reported no significant association. This study comprised about 150
    evaluated children, and would influence the overall evidence in a less
    significant direction if the data were able to be included.

    8.3.6.3  A quantitative assessment of the cross-sectional studies

         There have been more cross-sectional studies that have related
    body burden to full-scale IQ in school-age children, but they cannot
    all be included in a single meta-analysis. Here attention is focused
    on the PbB studies, and it would be appropriate subsequently to

    FIGURE 15

    FIGURE 16

    undertake an equivalent meta-analysis of the tooth lead studies. The
    studies included have differed in the number and nature of covariates.
    In particular, in most of the groups within the London and European
    multi-centre study (MCS), maternal IQ has not been controlled. Studies
    in which regression coefficients could not be obtained were excluded.
    Since the studies not included in the analysis account for only a
    small fraction of the total children investigated, their exclusion has
    a negligible effect on the meta-analysis estimates of effect size.

         The following studies are included in the meta-analysis:

         Lavrion, Greece               (Hatzakis et al., 1989)
         Edinburgh                     (Fulton et al., 1987)
         London (Greenwich)            (Yule et al., 1981)
         European multi centre study   (Winneke et al., 1990)

         The European multi-centre study comprises seven separate analyses
    for children in Bucharest, Budapest, Modena, Sofia, Dusseldorf (two
    studies) and Zagreb. The Lavrion centre in this study was separately
    reported by Hatzakis et al. (1987).

         As in the meta-analysis for the prospective studies, individual
    cross-sectional studies varied as to whether the logarithm of the PbB
    level or untransformed PbB level was used in the multiple regression
    analysis. Hence, as before, all associations are expressed in terms of
    the estimated changes in full-scale IQ for a change in PbB level from
    0.48 to 0.96 µmol/litre (10 to 20 µg/dl), after adjustment for
    confounders, as shown in Table 13.

         Fig. 17 shows these estimates and their 95% confidence limits for
    the 10 study samples. Only the two largest studies (Lavrion and
    Edinburgh) show statistically significant inverse associations, as
    indicated by confidence intervals entirely to the left of zero. The
    limited statistical power of the other studies is reflected in their
    wide confidence intervals. Despite the considerable variation in the
    study designs (e.g., sample selection, choice of confounders), there
    is no evidence of statistical heterogeneity. In other words, all the
    confidence intervals overlap and the heterogeneity test is not
    statistically significant.

         A combination of the evidence from all of these cross-sectional
    studies produces a more precise overall estimate of association, as
    shown in Fig. 17. This meta-analysis of the cross-sectional studies
    estimates that full-scale IQ is reduced by 2.15 points for an increase
    in PbB level from 0.48 to 0.96 µmol/litre (10 to 20 µg/dl), with a 95%
    confidence interval from -1.2 points to -3.1 points (P < 0.001).

    FIGURE 17

    8.3.6.4  Task group overview of cross-sectional studies

    a)   Methods of controlling for confounders

         A particular methodological concern in the cross-sectional
    studies is the manner in which they have taken account of confounding
    factors in deriving adjusted estimates of associations between body
    lead burden and neuropsychological performance, in particular IQ.

         The three aspects of the problem are:

    i) Which potential confounders are measured? There is considerable
    variation between studies here, suggesting that no individual study
    can have fully corrected for the full range of parental and social
    influences on IQ.

    ii) How were confounders selected for inclusion in the analysis? Some
    studies include confounders solely on the basis of their strength of
    association with outcome (IQ), while others have given attention to
    their association with body lead burden. It is difficult to assess the
    extent to which these differences would affect the results, but the
    former is more generally accepted statistical practice.

    iii) Which statistical analysis strategy was employed in reaching a
    final model? Most studies have adopted some form of multiple
    regression technique, but have varied as to whether forward selection
    (i.e. model building), backwards selection (i.e. model collapsing) or
    more arbitrary choices of final model were employed. In the larger
    studies, the choice between such strategies is unlikely to matter, but
    in smaller studies the greater play of random variation could make the
    results dependent on the statistical technique.

    b)   What they told us

         In most studies a negative association between lead measures and
    IQ measures is found in uncontrolled data. This difference is usually
    in the range of 4 to 6 IQ points. Most studies also confirmed the
    positive association between lead measures and indicators of social
    disadvantage, whether this is indicated by SES, maternal education or
    other more detailed indicators of non-optimal child-rearing
    environments, such as marital quality or maternal depression. When
    these social and other confounding factors are controlled, the effect
    has been, in most cases, to reduce the strength of the association
    between lead measures and IQ, although it remains in the same
    direction. Where maternal intelligence has not been controlled, the
    impact of correction for covariates tends to be smaller, and where
    more detailed social measures have been controlled, the impact has
    been greater.

    c)   What they did not tell us

         The cross-sectional nature of these studies, and the fact that
    many used a single measure of current exposure, limited their
    usefulness in answering questions relating to the natural history of
    the association between lead exposure and outcomes, including whether
    there were critical periods of exposure, and whether lead-associated
    deficits were persistent or reversible. Equally they were limited in
    the information that could be obtained relating to the natural history
    of confounders. They were also unable to answer questions of reverse
    causality.

    8.3.6.5  An interpretation of the overview of prospective and
             cross-sectional studies

         The above meta-analyses of the prospective studies and cross-
    sectional studies reveal a consistency between studies which points
    towards a "collectively significant" inverse association between PbB
    level and full-scale IQ. Taking the results in Fig. 15 as a guideline,
    there appears to be a mean decrease in full-scale IQ of the order of 2
    IQ points for a change in mean PbB level from 0.48 to 0.96 µmol/litre
    (10 to 20 µg/dl).

         Below this range, uncertainties are increased, concerning firstly
    the existence of an association and secondly estimates of the
    magnitude of any putative association. The relatively limited numbers
    of children in most studies with PbB levels below this range, the
    strong contributions of confounding variables, and limitations in the
    precision in analytical and psychometric measurements combine to lower
    statistical power to detect associations and to estimate their
    magnitude.

         The key question is whether this statistical association is
    directly attributable to the causal influence of lead on child IQ. It
    is important to consider alternative explanations as follows:

    a)  Chance The consistency of alternative analyses and the level of
    significance achieved suggest that chance cannot be a complete
    explanation. However, the confidence intervals are relatively wide, so
    that the magnitude of true association could be as low as a < 1-point
    (rather than a 2-point) deficit, or as high as > 3 points.

    b)  Confounding factors The adjusted lead/IQ associations tend to be
    substantially weaker than the unadjusted associations, which indicates
    that confounding factors are important. Since none of the studies can
    claim to have taken complete account of confounders, (e.g., father's
    IQ is not included), it seems likely that some of the remaining
    lead/IQ relationship, after adjustment, may still be attributable to a
    degree of unexplained confounding.

    c)  Reverse causality One initial justification for the prospective
    studies was that they could measure early exposure to lead and relate
    it to later child development, thus removing the problem of reverse
    causality potentially present in the cross-sectional studies. However,
    it has turned out that very early lead exposure, e.g., in the
    perinatal period, is not related to school-age IQ. The question arises
    as to whether children of lower IQ could exhibit behaviour patterns at
    earlier ages, e.g., around age 2 and older, which could enhance their
    uptake of lead. Perhaps, such reverse causality remains a possibility
    even in the prospective studies.

    d)  Selection biases The positive studies may perhaps be more likely
    to report in more quantitative details, as indicated by the absence of
    relevant data from Cleveland in the prospective studies. Post-hoc
    selection based on the more significant PbB concen-trations may lead
    to exaggerated estimates.

         It is a matter of debate and conjecture as to the extent to which
    these four issues should inhibit claims of a causal relationship in
    the prospective studies. The essential problem is that observational
    epidemiology cannot provide definitive evidence of causality when the
    key statistical association is weak, the temporal relationship is
    unclear and major confounders are present.

    8.3.7  Hearing impairment in children

         Schwartz & Otto (1987) reported that the probability of elevated
    pure-tone hearing thresholds at 500, 2000 and 4000 Hz increased
    significantly with increasing PbB level in 4519 subjects (4-19 years
    of age) who participated in the NHANES II study. Variables included in
    the backwards stepwise regression models included information derived
    from medical history, clinical examinations, and assessment of
    developmental milestones.

         Schwartz & Otto (1991) examined data for subjects (6-19 years of
    age) from the Hispanic Health and Nutrition Examination Survey
    (HHANES) study which included three distinct ethnic groups. After
    excluding 283 subjects with previous ear problems (discharges,
    ruptured ear drums or tinnitus), data for 3262 were analysed. The
    authors concluded that increasing PbB level in the range
    0.36-0.86 µmol/litre (7-18 µg/dl) was associated with approximately
    2 dB loss of pure-tone hearing at frequencies of 500, 1000, 2000 and
    4000 Hz.

         Dietrich et al. (1992) assessed the relationship between scores
    on a test of sensory auditory processing (SCAN) and prenatal/postnatal
    PbB concentrations in 215 subjects drawn from the Cincinnati
    prospective cohort study. Higher prenatal, neonatal and postnatal PbB
    concentrations were associated with more incorrect identification of
    common monosyllabic words presented under conditions of muffling.

    Other variables associated with impaired central auditory processing
    were pure-tone audiometry results, social class, quality of caretaking
    in the home, birth weight, gestational age, a measure of obstetrical
    complications, and consumption of alcohol during pregnancy. Following
    adjustment for these co-factors, lifetime average PbB concentration
    remained significantly and inversely associated with SCAN performance.

    8.4  Renal system

         Acute exposure to lead is known to cause proximal renal tubular
    damage, characterized by generalized aminoaciduria, hypophosphataemia
    with relative hyperphosphaturia, and glycosuria (Chisolm, 1962).
    Cellular structural changes include nuclear inclusion bodies,
    mitochondrial changes and cytomegaly of the proximal tubular
    epithelial cells (Cramer et al., 1974). Diagnosis of lead-induced
    altered renal function or disease is difficult since there are no
    specific indicators; blood urea nitrogen (BUN) and creatinine levels
    become elevated only when twothirds of renal function has been lost
    (Bernard & Becker, 1988).

    8.4.1  Clinical studies

         From a study of seven men occupationally exposed to lead in a
    shipyard during oxy-acetylene flame cutting of lead-painted steel
    hulls, Cramer et al. (1974) concluded that there is a continuum of
    morphological and functional change in the pathogenesis of chronic
    lead nephrotoxicity. Each subject was treated in hospital 3 or more
    days after recent exposure and PbB levels exceeded 3.36 µmol/litre
    (70 µg/dl) in all cases. On the basis of microscopic examination of
    biopsy tissue, nuclear inclusion bodies were reported in proximal
    tubular cells during the early phase but there was no impairment of
    renal function. In a second phase, there was fibrosis associated with
    asymptomatic azotaemia and reduced glomerular filtration rate but
    without demonstrable proximal tubular dysfunction; renal failure was
    not seen in the study.

         Weeden et al. (1979) diagnosed lead nephropathy in 15 lead
    workers, all having reduced glomerular filtration rates. Renal
    biopsies of six of the subjects showed focal interstitial nephritis in
    addition to non-specific changes in the proximal tubules. At the time
    of examination, PbB levels for 11 of the 15 workers were within the
    range 1.92-3.84 µmol/litre (40-80 µg/dl).

         Baker et al. (1979) reported increased BUN and decreased
    creatinine clearance in 28 workers, all of whom had relatively
    prolonged, high-dose lead exposure in lead smelting or chemical
    manufacturing.

         A study by Maranelli & Apostoli (1987) of 60 workers, described
    as "lead poisoned" and having PbB levels of 3.45 ± 0.80 µmol/litre
    (71.9 ± 16.6 µg/dl), found no definitive correlation between PbB, lead
    in urine after chelation, and BUN, serum creatinine and serum uric
    acid.

    8.4.2  Epidemiological studies

    8.4.2.1  Occupational cohorts

         Several recent studies of lead-exposed workers provide further
    information relating to dose-effect relationships. The emphasis in
    these studies was on lead and its possible association with adverse
    effects on health. However, exposure to other potentially toxic
    substances in the working and living environment should not be
    overlooked/ignored.

         For example, Buchet et al. (1980) examined 25 male lead smelter
    workers and 88 male control workers. The PbB levels of the lead
    workers were in the range 1.62-2.94 µmol/litre (33.8-61.3 µg/dl) for a
    mean of 13.2 years (range 3.1-29.8 years) of lead exposure. The PbB
    levels for the controls were in the range 0.26-1.64 µmol/litre
    (5.5-34.2 µg/dl). There were no differences for parameters of renal
    function between the groups and no signs of clinical renal impairment.
    It was concluded that a PbB level of less than 2.98 µmol/litre
    (62 µg/dl) is not associated with renal toxicity.

         N-acetyl-ß-d-glucosaminidase (NAG) is a lysosomal enzyme present
    in renal tubular cells. This enzyme is a sensitive but non-specific
    indicator for early sub-clinical nephrotoxicity. In a study on 29
    lead-exposed workers, Meyer et al. (1984) found increased NAG in
    urine, but there was no correlation with PbB level. However, NAG level
    was found to be normal in the 5 subjects with PbB levels exceeded
    3.36 µmol/litre. These authors speculated that long-term high exposure
    to lead may deplete the kidneys of NAG or render it insensitive to the
    effects of lead exposure.

         Verschoor et al. (1987) studied 155 male lead workers and 126
    control workers who were matched for age, smoking habits,
    socioeconomic status and duration of employment. The PbB levels of the
    lead workers were in the range of 0.43-4.71 µmol/litre
    (8.3-97.6 µg/dl) compared with the controls 0.15-0.96 µmol/litre
    (3.1-18.8 µg/dl). The lead workers had elevated ZPP levels
    (34-292 µmol/mol haemoglobin) compared with controls (10-35 µmol/mol
    haemoglobin). No significant differences were found for various
    indicators of renal function; all urinary and serum parameters were
    within normal ranges. There were no differences in protein excretion
    patterns and no signs of renal impairment. However, the authors found
    that NAG levels in the lead-exposed workers were higher than control
    values and increased with increasing PbB levels. They concluded that
    lead exposure resulting in PbB levels of under 3.0 µmol/litre

    (62 µg/dl) can affect renal tubular functions as measured by NAG
    excretion; lead appeared to affect the tubular parameters more than
    the glomerular parameters in moderately exposed workers.

         Ong et al. (1987) examined 158 male and 51 female lead battery or
    smelter workers and 30 control workers. The lead workers had 1-36
    years exposure with an average of 10.8 ± 8.0 years and PbB levels in
    the range of 0.14-3.84 µmol/litre (3-80 µg/dl); only five workers
    exceed 2.88 µmol/litre (60 µg/dl). The authors found a weak but
    statistically significant positive association between PbB and blood
    urea nitrogen, and between PbB and serum creatinine, and that
    creatinine clearance was reduced with increasing PbB level. NAG levels
    in the lead-exposed workers were significantly higher than control
    values and increased with increasing urine lead level when the data
    were adjusted for age. These authors concluded that a relatively low
    PbB level can affect renal function.

         Not all investigators have found parallel association between PbB
    and NAG levels. Gennart et al. (1992b) compared 98 lead-exposed lead
    acid battery workers (mean PbB level 2.45 µmol/litre, 51 µg/dl,
    geometric mean ZPP 10.2 µg/g Hb) with 85 controls (mean PbB level
    1.00 µmol/litre, 20.9 µg/dl, geometric mean ZPP 2.84 µg/g Hb)
    recruited from other departments in the same factory. None of the
    indicators of renal function (retinol-binding protein,
    ß2-microglobulin, albumin or NAG in urine, or creatinine or
    ß2-microglobulin in serum) were correlated with PbB level, duration
    of exposure or ZPP, or showed significantly different mean values
    between the lead-exposed and control groups of workers.

         Cardenas et al. (1993) recently evaluated 27 different
    indications of renal dysfunction in 50 workers exposed to lead and 50
    controls. There were significant increases in urinary NAG and sialic
    acid in the lead exposed group. These changes may represent minor
    cellular modifications rather than significant functional or
    irreversible renal damage. There was a significant decrease in urinary
    6-keto PGF1alpha and a significant increase in TXB2. These
    eicosanoid changes may reflect systemic functional vascular changes
    rather than an effect of lead on the kidney.

    8.4.2.2  General population

         An epidemiological survey on 283 persons in Scotland, from
    households with water lead concentrations in excess of 100 µg/litre,
    revealed a close correlation between water lead content and PbB and
    serum urea concentrations (Campbell et al., 1977). The frequency of
    renal dysfunction in individuals with elevated PbB concentrations
    (> 2 µmol/litre or > 41 µg/dl) was significantly greater than that
    of age- and sex-matched controls.

         Pocock et al. (1984) measured serum creatinine, urate and urea
    concentrations in 7364 British men, and 74 subjects had PbB levels
    equal to or greater than 1.8 µmol/litre (37.3 µg/dl). The authors
    concluded that there was no indication that exposure to lead at
    concentrations commonly encountered in Britain was responsible for
    impaired renal function.

         Staessen et al. (1992) investigated the relationship between lead
    exposure and renal function as part of a cross-sectional population
    study of the health effects of environmental exposure to cadmium.
    Creatinine clearance, PbB and ZPP were measured in a random population
    sample of 965 men (geometric mean PbB level was 0.55 µmol/litre or
    11.4 µg/dl) and 1016 women (geometric mean PbB level was
    0.36 µmol/litre or 7.5 µg/dl). Creatinine clearance rate (mean was
    99 ml/min in men, 80 ml/min in women) was inversely correlated with
    PbB and ZPP levels before and after adjustment for age, body mass
    index and diuretic treatment. Also positively correlated were serum
    ß2-microglobulin and PbB level in men, serum ß2-microglobulin and ZPP
    in men and women, and serum creatinine and ZPP in men. Impaired renal
    function could not be explained by exposure to cadmium or by elevated
    blood pressure. The authors concluded that exposure to lead may impair
    renal function in the general population. However, the study could not
    exclude the possibility that renal impairment may lead to an increase
    in PbB level.

    8.4.2.3  Cohort mortality studies

         "Lead poisoning" was a diagnosis given to 241 workers employed
    for 1-30 years at the Port Pirie lead smelter between 1928-1959 by a
    State medical board. Death registration records for the period
    1930-1977 identified 140 deaths among the group. The cause of death
    profile was compared with that of 695 other male decedents,
    predominantly production workers and a smaller number of office
    workers. Age-standardized mortality analysis revealed a substantial
    excess of deaths attributed to chronic nephritis and to cerebral
    haemorrhage. The rates for lead-poisoned workers exceeded those for
    the other workers, and both exceeded the rates for the Australian
    general male population. The rates decreased over successive calendar
    periods but excess rates persisted for chronic nephritis up until 1977
    (McMichael & Johnson, 1982).

    8.5  Cardiovascular system

         Two persistent issues have been under intense study since
    publication of Environmental Health Criteria 3: Lead (IPCS, 1977):

    a)   whether lead is a factor in hypertension and, if so, whether
         there is a causal relationship;

    b)   whether lead is contributory to cardiovascular effects
         influencing morbidity or mortality.

    8.5.1  Blood pressure

    8.5.1.1  Studies on occupationally exposed cohorts

         From a study of 431 white male police officers 24-55 years of age
    (Moreau et al., 1982; Orssaud et al., 1985), it was concluded that
    systolic and diastolic blood pressure was related to PbB level, the
    correlation being greatest for the younger subjects and decreasing
    with age. Statistical adjustment was made for alcohol consumption and
    body mass index, but not for smoking. The magnitude of observed
    association between systolic pressure and PbB level in this small
    study was somewhat greater than in the larger general population
    studies cited below.

         Parkinson et al. (1987) studied 270 lead battery workers and 158
    non-exposed workers and compared lead exposure with systolic and
    diastolic blood pressures. After controlling for age, education,
    income, cigarette smoking, alcohol consumption and exercise, there was
    a small and non-significant association. The average PbB level of
    lead-exposed workers was 1.92 ± 0.62 µmol/litre (40 ± 13 µg/dl),
    whereas in non-exposed workers it was 0.34 ± 0.24 µmol/litre
    (7 ± 5 µg/dl).

    8.5.1.2  Studies in the general population

         The possible relationship between PbB concentration and blood
    pressure has been examined in several large-scale population studies.
    These include the British Regional Heart Study (BRHS), the US NHANES
    II (National Health and Nutrition Examination Survey) and studies in
    Wales, Denmark, Canada and Belgium.

         The BRHS is a prospective study of 7735 men, initially aged 40-59
    years, from 24 British towns, who were first examined in 1978-1980
    (Shaper et al., 1981). The median PbB level was 0.7 µmol/litre
    (14.5 µg/dl). The initial findings (Pocock et al., 1984), which
    indicated a lack of association between PbB level and blood pressure,
    were later re-examined to take account of potential confounding
    factors such as alcohol, smoking and town of residence. Pocock et al.
    (1988) noted that 95% confidence intervals for systolic and diastolic
    blood pressure plotted against PbB concentrations overlapped and
    showed no elevation even at the highest observed PbB concentration.
    Applying statistical techniques to adjust mean systolic and diastolic
    blood pressure measurements for body mass index, age, alcohol
    consumption, cigarette smoking and town of residence showed signs of a
    weak association, particularly for diastolic pressure and PbB level.
    Multiple regression analysis showed that adjustment for personal
    characteristics but not for town of residence rendered both systolic
    and diastolic regressions on log PbB insignificant, chiefly because
    alcohol consumption is an important confounder, being positively
    related to both blood pressure and PbB level. Introducing an extra
    adjustment for town of residence made the systolic and diastolic

    regressions highly statistically significant, even though the
    associations were weak. These cross-sectional data indicated that for
    every doubling in PbB level (e.g., from 0.8 to 1.6 µmol/litre) there
    are estimated mean increases of 1.45 mmHg (95% confidence interval
    0.47-2.43 mmHg) in systolic blood pressure and 1.25 mmHg (95%
    confidence interval 0.65-1.85 mmHg) in diastolic blood pressure.

         The NHANES II study, a USA national cross-sectional survey
    carried out in 1976-1980, included PbB and blood pressure measurements
    in a general population sample of 5803 men and women aged 12-74.
    Geometric mean PbB levels were around 0.72 µmol/litre (about 15 µg/dl)
    for men and approximately 0.53 µmol/litre (11 µg/dl) for women.
    Several authors have examined these data for possible PbB and blood
    pressure associations.

         In an analysis controlling for the confounding factors of age,
    race and body mass index, Harlan et al. (1985) and Harlan (1988) found
    a significant association between PbB level and blood pressure for men
    but not for women. White males aged 40-59 years from the same data set
    were analysed by Pirkle et al. (1985), and the correlation between PbB
    level and blood pressure was confirmed. In further analysis of the
    whole age range, adjustments for possible time trend and geographical
    site effects did not affect the significance of the association in
    males (Landis & Flegal, 1988; Schwartz, 1988), although its magnitude
    did become somewhat less pronounced.

         Gartside (1988) applied a method of forward stepwise regression
    to PbB and blood pressure data from NHANES II and reported that the
    results for white men, white women and black men were contradictory
    and lacked consistency and reliability. He noted that the overall
    average was too small to conclude association between PbB level and
    blood pressure in this study.

         Two surveys were carried out in Wales by Elwood et al. (1988a,b).
    The Welsh Heart Programme carried out in 1985 provided complete PbB,
    blood pressure and other data for 865 men and 856 women aged 18-64.
    The geometric mean PbB level for men was 0.56 µmol/litre (11.6 µg/dl)
    and for women was 0.43 µmol/litre (9.0 µg/dl). In neither sex was
    there a significant correlation between PbB level and systolic or
    diastolic blood pressure. The second study was in a cohort of men aged
    49-65 years living in Caerphilly, Wales. The geometric mean PbB level
    (N = 1137) was 0.61 µmol/litre (12.7 µg/dl). Complete data from 1137
    subjects, ranking blood pressure readings according to PbB level, did
    not reveal any trend in the percentage of subjects with systolic
    pressure above 160 mmHg. The authors corrected only for age as a
    confounding factor.

         Data from a Canadian study (Neri et al., 1988) collected during
    10 months in 1978-1979 for 2193 subjects aged 25-64 showed a weak but
    statistically significant association between PbB level and diastolic
    blood pressure.

         Grandjean et al. (1989b) studied 504 men and 548 women residing
    in Glostrup (Denmark) at age 40; 451 men and 410 of the women were
    followed-up five years later. Average PbB levels for the men at 40 and
    45 years of age were 0.62 µmol/litre (13 µg/dl) and 0.43 µmol/litre
    (9 µg/dl), and for the women 0.43 µmol/litre (9 µg/dl) and
    0.29 µmol/litre (6 µg/dl), respectively. All correlations found
    between PbB and blood pressure became insignificant when blood
    haemoglobin and alcohol intake were entered into multiple regression
    analyses; an independent effect of low-level lead exposure on blood
    pressure could not be distinguished.

         Möller & Kristensen (1992) examined 1052 men and women in
    Copenhagen, Denmark, in 1976, and re-examined them in 1981 and 1987
    (men only). Initial mean PbB levels were 0.65 µmol/litre (13.6 µg/dl)
    in men and 0.46 µmol/litre (9.6 µg/dl) in women, and had fallen by
    around 30% 5 years later. For men, there was no significant
    association between PbB level and blood pressure (or between changes
    in PbB level and blood pressure) after adjustment for confounders
    (tobacco, alcohol, body mass and physical activity). For women, the
    association between PbB level and diastolic (but not systolic) blood
    pressure remained significant on both occasions, even after
    controlling for confounders.

         A Belgium study (Dolenc et al., 1993) included 827 men and 821
    women (mean age 25) whose mean PbB levels were 0.56 and
    0.34 µmol/litre (11.6 and 7.07 µg/dl), respectively. After adjustment
    for covariates (body mass index), pulse rate and serum creatinine and
    serum calcium levels), systolic blood pressure was inversely
    associated with PbB level in men (P < 0.05). There were no
    significant PbB associations for diastolic blood pressure in men or
    for either pressure in women.

         A study of 398 male and 133 female civil servants (Staessen et
    al., 1990) showed geometric mean PbB levels of 0.58 µmol/litre
    (12.06 µg/dl) and 0.46 µmol/litre (9.56 µg/dl) in men and women,
    respectively. Taking the data for both sexes together, there were
    statistically significant positive associations between PbB level and
    systolic and diastolic blood pressures (r = +0.11 in each case) but
    these became non-significant after adjustment for confounders (sex,
    age, BMI, pulse, delta-glutamyltranspeptidase (delta-GTP) and serum
    calcium level).

         Another cross-sectional study involving only women was performed
    in Boston (Rabinowitz et al., 1987). Cord PbB levels (mean
    0.33 µmol/litre, 6.9 µg/dl) among 3851 women correlated with both the
    systolic (r = 0.08) and diastolic (r = 0.05) blood pressures measured
    during and before delivery. Multivariate models of pregnancy
    hypertension as a function of age, parity, haematocrit, diabetes,
    ponderal index, and race were improved when lead was included as a

    predictor. Lead appeared to have a small but demonstrable association
    with pregnancy hypertension and blood pressure at delivery, but not
    with eclampsia.

         In an overview of the BRHS, NHANES II and Welsh studies, Pocock
    et al. (1988) concluded that with overlapping confidence limits the
    data provided weak but reasonably consistent evidence of lead and
    blood pressure associations. They noted that the NHANES II data on
    2254 men in the USA indicate a slightly stronger association between
    PbB level and systolic blood pressure, whereas data from over 2000 men
    in Wales did not show a statistically significant association. They
    inferred that a causal relationship could not be concluded from any of
    the epidemiological studies.

         Staessen et al. (1994) have undertaken a more extensive
    meta-analysis of nearly all of the above studies and some other
    smaller ones as well. This amounted to a total of 19 studies with
    28 210 subjects. They found that the association between PbB level and
    blood pressure was similar in both sexes. In all studies combined, a
    two-fold increase in PbB concentration was associated with a 1.0 mmHg
    increase in systolic pressure (95% confidence interval (CI); 0.3 to
    1.7 mmHg; P = 0.008) and a 0.7 mmHg increase in diastolic pressure
    (95% CI: 0.2 to 1.3 mmHg; P = 0.02). These authors conclude that the
    published evidence suggests a weak positive association between blood
    pressure and lead exposure, but any such relationship may not be
    causal and is unlikely to entail any public health implications
    regarding hypertension.

    8.5.2  Other cardiovascular effects

         As recorded previously (IPCS, 1977), there is good evidence that
    signs of clinical lead poisoning sometimes include evidence of toxic
    action on the heart. Kopp et al. (1988) reviewed the cardiovascular
    actions of lead and concluded that the degree of cardiovascular
    involvement during episodes of acute lead intoxication depends on the
    duration of exposure and dose. It is not known whether environmental
    exposure to lead affects the electrical or mechanical activity of the
    heart.

    8.5.2.1  Occupational studies

         There have been several studies relating to cardiovascular
    effects in lead-exposed occupational groups. However, as with any
    study of an occupational cohort, it is important to evaluate the
    strength of the cause-effect relationship, the nature or substance of
    other contributory factors and how well the workplace "exposure" has
    been evaluated for other potential toxicants. Thus, the occupational
    hazards associated with work in lead smelting or lead battery
    operations include, but are not limited to, lead.

    8.5.2.2  Studies in the general population

         Pocock et al. (1988) found that after 6 years of follow-up of the
    BRHS cohort of 7735 middle-aged men, 316 of the men had major
    ischaemic heart disease and 66 had had a stroke. After allowance for
    confounding effects of cigarette smoking and town of residence, there
    was no evidence that PbB was a risk factor for these cardiovascular
    events.

         Möller & Kristensen (1992) studied 1052 men and women in
    Copenhagen, Denmark, and related blood pressure to subsequent
    cardiovascular morbidity and mortality over a 14-year period. There
    were significant positive associations between PbB level and both
    coronary and cardiovascular disease in univariate analysis, but these
    became non-significant after controlling for confounders.

    a)   Clinical studies

         Boscolo & Carmignani (1988) measured plasma renin activity in
    hospitalized lead-exposed workers and concluded that synthesis or
    release of renin is increased by short or moderate exposure to lead
    and decreased with prolonged exposure.

    b)   Epidemiological studies

         Kirkby & Gyntelberg (1985) found a significantly higher incidence
    of ischaemic ECG changes in a study of lead smelter workers (20%, mean
    PbB level 2.45 µmol/litre, 51 µg/dl), compared with matched
    non-exposed controls (6%, mean PbB level 0.53 µmol/litre, 11 µg/dl).
    There was also a slight (4-5 mmHg) increase in diastolic blood
    pressure in the lead workers compared with the controls.

    c)   Cohort mortality studies

         Two studies (Cooper, 1988 and Fanning, 1988) reported an
    increased mortality rate of lead workers due to circulatory disease,
    but no such relationship was found in reports of studies by
    Gerhardsson et al. (1986) or Selevan et al. (1988). Gerhardsson et al.
    (1986) examined a lead-exposed sub-cohort of 437 from a study cohort
    of 3832 male workers first employed at a copper smelter before 1967
    and followed up from 1950 to 1981. There was no excess mortality due
    to ischaemic heart disease or cerebrovascular disease in the
    lead-exposed worker sub-cohort. Selevan et al. (1988) found no
    association between lead exposure and deaths due to hypertensive
    diseases in a review of mortality in a cohort of male hourly workers
    at a lead smelter. There were many confounding factors in each of
    these studies.

    8.5.3  Summary

         Despite intensive efforts to define the relationship between body
    burden of lead and blood pressure or other effects on the
    cardiovascular system, no causal relationship has been demonstrated in
    humans and the mechanisms remain obscure.

         There is experimental evidence from animal studies indicative of
    an effect of lead on blood pressure, and several mechanisms have been
    proposed to explain these observations (see section 7.4).

    8.6  Gastrointestinal effects

    8.6.1  Occupational exposure

         Colic is a well-recognized symptom of acute lead poisoning and is
    still reported in groups of lead-exposed industrial workers. Of
    particular concern are the reports of intoxication caused by acute
    exposure (inadequate protective measures) to lead associated with
    removal of lead-based paint by burning or sand-blasting and the
    demolition of lead-containing industrial plants. Symptoms of colic
    include abdominal pain, constipation, cramps, nausea, vomiting,
    anorexia, weight loss and decreased appetite.

         Symptoms in adults typically occur at PbB levels of
    4.8-9.6 µmol/litre (100-200 µg/dl) but have been noted at levels as
    low as 1.92 µmol/litre (40 µg/dl) (Baker et al., 1979; Haenninen et
    al., 1979; Awad El Karim et al., 1986; Pollock & Ibels, 1986; Muijser
    et al., 1987; Holness & Nethercott, 1988; Marino et al., 1989;
    Pagliuca et al., 1990; Schneitzer et al., 1990).

         Positive histories of lead colic were given by 40 workers in a
    cohort of 158 secondary lead smelter workers who participated in a
    clinical field survey (Lilis et al., 1977). It was reported most
    frequently by those having PbB levels above 3.84 µmol/litre (80 µg/dl)
    at the time of examination and was not found where ZPP was within the
    normal range. Thirty percent of workers having ZPP > 200 µg/g Hb had
    experienced symptoms.

         In a study of a population of 585 black South African factory
    workers, Irwig et al. (1978) reported that the incidence of abdominal
    pain increased with increasing PbB level: it was 12% for the first
    quartile (PbB < 3.28 µmol/litre or < 68 µg/dl), 23% for the
    second (3.28-4.14 µmol/litre or 68-86 µg/dl), 24% for the third
    (4.15-5.14 µmol/litre or 86-107 µg/dl) and 37% for the fourth
    (> 5.15 µmol/litre or > 107 µg/dl).

         In a cross-sectional clinical study of 90 telephone
    cable-splicers by Fischbein et al. (1980), 19 (21%) workers reported
    gastrointestinal symptoms (mean PbB level of 1.44 µmol/litre, or

    30 µg/dl and mean ZPP of 66.6 µg/g Hb) whereas 71 (79%) reported no
    such symptoms (mean PbB level of 1.3 µmol/litre or 27 µg/dl and mean
    ZPP of 52.3 µg/g Hb).

    8.6.2  Exposure of children

         Colic is seen in children and US EPA (1986a) concluded that the
    lowest-observed-adverse-effect level was in the range of
    2.88-4.80 µmol/litre (60-100 µg/dl).

    8.7  Liver

         There appears to be no new evidence relating human body lead
    burden to effects on the liver, but it has been suggested that the
    effects of lead on haem synthesis may alter the functional capacity of
    the hepatic cytochrome P-450 system to metabolize drugs.

    8.7.1  Occupational exposure

         Fischbein et al. (1977) tested five demolition workers acutely
    exposed to lead for a period of 3 months prior to the study. They
    found that the plasma half-life of an oral dose of antipyrine was
    within the range for normal healthy volunteers; it was shorter in each
    subject after chelation therapy but still within the normal range. The
    plasma half-lives of phenylbutazone were also within the normal range
    but were unaffected by chelation therapy.

    8.7.2  Exposure of children

         Saenger et al. (1984) found decreased urinary excretion of
    6-ß-hydroxycortisol in 26 children with a mean PbB level of
    2.11 µmol/litre (44 µg/dl); the decreased formation of the metabolite
    was attributed to lead inhibition of the cytochrome P-450-dependent
    mixed-function oxidases.

    8.8  Reproduction

         While it is generally accepted from early literature that lead
    adversely affects the reproductive process in both men and women, the
    evidence is mostly qualitative and dose-effect relationships have not
    been established.

         Most new information relates to reports of occupational cohorts
    and of populations living in polluted areas near industrial plants.
    There is qualitative evidence that lead is toxic to the reproduction
    system in both men and women. However, there are insufficient data to
    provide the basis for estimation of dose-effect relationships in
    women.

    8.8.1  Female populations

         Nordstrom et al. (1978b) reported an increased frequency of
    spontaneous abortion in women living close to a smelter in northern
    Sweden. In a later report, Nordstrom et al. (1979) described the
    responses to a questionnaire completed by 511 of 662 women who had
    worked at the smelter and were born between 1930 and 1959. Spontaneous
    abortion rates were highest in those pregnancies in which the mother
    was employed during the pregnancy (13.9%) or had been employed before
    the pregnancy and was living close to the smelter (17%); the frequency
    rate was higher (19.4%) when the father worked at the smelter.
    However, it should be noted that the smelter produced copper and lead
    in addition to a number of other metallurgical and chemical products
    (Nordstrom et al., 1978a) and that the effects reported may not
    necessarily be attributable exclusively to lead.

         A study of pregnancies in the centre and surrounding areas of the
    lead smelter town of Port Pirie showed that the incidence of
    miscarriages (22 or 23) and stillbirths (10 or 11) was higher in women
    living close to the smelter (McMichael et al., 1986). However, in a
    study of 639 Yugoslav women, the risk of spontaneous abortion was not
    increased (OR = 1.1, 95% CI: 0.9-1.4) among women with mean PbB levels
    of 0.77 µmol/litre (16.0 µg/dl) as compared with women with mean PbB
    levels of 0.25 µmol/litre (5.2 µg/dl) (Murphy et al., 1990). Risk did
    not vary with distance of residence from the smelter.

         Some studies have found decreased length of gestation in women
    whose PbB levels were greater than 1.09 µmol/litre (23 µg/dl) (Moore
    et al., 1982), 0.58 µmol/litre (12 µg/dl) (Dietrich et al., 1986) or
    0.72 µmol/litre (15 µg/dl) (McMichael et al., 1986). However, neither
    Bellinger et al. (1991b) nor Graziano et al. (1990) found decreases in
    gestational length or other parameters of pregnancy in women with
    elevated PbB levels.

         In the Cincinnati prospective study, a significant reduction in
    birth weight associated with prenatal (maternal) PbB levels, after
    adjustment for covariates, was reported by Bornschein et al. (1989).
    In the Port Pirie prospective study of 749 pregnancies (McMichael et
    al., 1986), the proportion of pregnancies resulting in low birth
    weight singleton infants was more than twice as high in Port Pirie
    women (whose PbB levels averaged 0.50 µmol/litre, 10.4 µg/dl) than in
    women outside Port Pirie (average PbB level of 0.264 µmol/litre,
    5.5 µg/dl). On the other hand, multiple regression analyses showed no
    significant association between low birth weight and maternal PbB
    level. In a cross-sectional study, Ward et al. (1987) reported a
    significant simple relationship between placental lead concentrations
    and reduced birth weight and head circumference.

         Other studies have not shown a significant association between
    birth weight and lead exposure. The Kosovo prospective study failed to
    detect any evidence of lead-related birth weight reduction in more

    than 900 births (Murphy et al., 1990), and Ernhart et al. (1986) found
    no significant effect of lead on birth weight, birth length or head
    circumference.

    8.8.2  Male populations

         Reproductive effects from occupational exposure to lead include
    asthenospermia, hypospermia, teratospermia and hypogonadism (US EPA,
    1986a; Braunstein et al., 1987).

         In men, effects on sperm or the testes may result from chronic
    exposure to lead at blood levels of the order of 1.92-2.4 µmol/litre
    (40-50 µg/dl) (ATSDR, 1991).

         Wildt et al. (1983) compared two groups of lead-exposed storage
    battery workers and concluded that the highly exposed group (PbB of
    more than 2.16 µmol/litre, 45 µg/dl) had decreased prostate/seminal
    vesicle function, low semen volume and poorer functional maturity of
    sperm.

         Chowdhury et al. (1986) reported a significant decrease in sperm
    count and motility and an increased count of abnormal spermatozoa in
    lead-exposed men (average PbB level of 2.04 µmol per litre or
    42.5 µg/dl) compared with controls (0.71 µmol/litre or 14.8 µg/dl).
    Assenato et al. (1987) also reported decreased sperm production in 18
    battery factory workers having PbB levels of 2.4-2.93 µmol/litre
    (50-61 µg/dl).

         In five out of six symptomatic lead intoxicated workers having
    PbB levels in the range of 1.87-4.7 µmol/litre (39-98 µg/dl) at the
    time of examination, Cullen et al. (1984) found defects of
    spermatogenesis, including oligospermia and azoospermia.

         The results of Lerda (1992) also showed significant increases in
    asthenospermia and teratospermia among 38 battery workers with PbB
    levels of 1.95-4.7 µmol/litre (40.5-98.0 µg/dl) compared to a control
    group with PbB levels of 0.86-1.25 µmol/litre (17.6-26.0 µg/dl).

         An epidemiological study by Lindholm et al. (1991) of 213 wives
    of lead workers and 300 matched controls suggested an association
    between paternal lead exposure and the risk of spontaneous abortion
    among the wives of lead workers when the PbB level exceeded
    1.5 µmol/litre (31 µg/dl) close to the time of spermatogenesis.
    However, the authors acknowledged that exposure of the study subjects
    to other metals and organic solvents may have influenced the results
    of the study.

         A weak association between paternal exposure, estimated from
    occupational history, and risk of esotropia (inward deviation
    strabismas) was noted in a case control study of 377 children aged 7
    years (OR - 2.4, 95%, CI: 1.0-5.6) (Hakim et al., 1991). However, the

    increased risk was not dose-related, and the risk of exotropia
    (outward deviation) was not associated with higher paternal lead
    exposure.

    8.8.3  Hormonal responses

         Assenato et al. (1987) found no significant differences in FSH,
    testosterone, prolactin, LH and total neutral 17-ketosteroids levels
    between 18 lead battery workers (PbB level, 2.4-3 µmol/litre or
    50-61 µg/dl) and 18 cement workers (PbB level, 0.9-1.1 µmol/litre or
    18-22 µg/dl), despite the finding of oligospermia in the lead-exposed
    group.

         Rodamilans et al. (1988) studied the endocrine status of 23 lead
    smelters in relation to the duration of lead exposure. Five workers
    exposed for less than 1 year (mean PbB level, 0.29 µmol/litre or
    66 µg/dl) showed an increase in serum LH level, while that of
    testosterone remained normal. Eight workers exposed for 1-5 years
    (mean PbB level, 3.50 µmol/litre or 73 µg/dl) and 10 employed for more
    than 5 years (mean PbB level, 3.65 µmol/litre or 76 µg/dl) showed an
    increase in LH comparable with the group exposed for less than 1 year
    but there was a clear reduction in serum testosterone levels.

         Gustafson et al. (1989) found a lower plasma FSH level in 25
    moderately exposed workers (mean PbB level, 1.82 µmol/litre or
    38 µg/dl) than in 25 matched controls. A lower mean serum LH level was
    reported by McGregor & Mason (1990) in 90 lead workers (mean PbB
    level, 2.20 µmol/litre or 46 µg/dl) than in controls; there was also a
    correlation between serum FSH and PbB levels. In an epidemiological
    study involving 122 current lead workers (PbB mean level,
    1.69 µmol/litre or 35.1 µg/dl) and 49 non-exposed workers (PbB mean
    level, 0.4 µmol/litre or 8.3 µg/dl), Ng et al. (1991) found slightly
    higher plasma LH and FSH levels among the exposed workers. The
    testosterone level, however, was not significantly different between
    the two groups. In contrast, Gennart et al. (1992) did not find an
    effect of lead on various endocrine parameters (including FSH and LH
    levels) in a study population of 221 workers having a geometric mean
    PbB level of 2.45 µmol/litre (51.0 µg/dl). They suggested that the
    hypothalamic-pituitary system may not be influenced by moderate
    exposure to lead.

    8.8.4  Postnatal growth and stature

         Several reports have suggested that the physical growth and
    stature of children may be reduced by exposure to lead (e.g., Mooty et
    al., 1975; Johnson & Tenuta, 1979; Routh et al., 1979), but the
    influence of other factors (e.g., race and diet) has often made it
    difficult to isolate lead as a causal agent for such effects in human
    populations. Multivariate regression analyses of NHANES data for
    approximately 2700 children in the USA (Schwartz et al., 1986)
    provided more convincing evidence of a significant association between

    increasing PbB levels and reduced height, weight, and chest
    circumference after adjusting for age, race, sex and nutritional
    covariates.

    8.9  Effects on chromosomes

         As noted in Environmental Health Criteria 3: Lead (IPCS, 1977),
    the literature remains controversial concerning induction of
    chromosomal changes in human lymphocytes by  in vivo exposure to
    lead. Most of the studies are of small numbers of subjects in
    occupational groups where lead was one of many potentially toxic
    agents. However, some of the difficulties arise because of the lack of
    standard procedures used for the assessment of chromosomal effects in
    cultured lymphocytes and from a lack of understanding of the health
    significance of chromosomal abnormalities.

         Induced mitotic activity in peripheral lymphocytes and increases
    in the rates of abnormal mitosis were reported by Sarto et al. (1978).
    PbB levels in the range of 1.05-4.27 µmol/litre (22-89 µg/dl) were
    reported to be associated with increased incidence of chromosomal
    aberrations by Schmid et al. (1972) in 32 lead manufacturing workers,
    by Nordenson et al. (1978) in 18 smelter workers and by Al-Hakkak et
    al. (1986) in 19 lead manufacturing workers. It should be noted that
    exposure of worker to potentially hazardous agents other than lead was
    not documented in these studies.

         Maki-Paakkanen et al. (1981) reported no change in chromosomal
    aberration frequency among workers with PbB levels in the range of
    1.82-5.76 µmol/litre (38-120 µg/dl); weak effects were found by Huang
    et al. (1988b) when the PbB level exceeded 2.5 µmol/litre (52 µg/dl).

         Qazi et al. (1980) reported increased numbers of cells with
    chromosome breaks in a baby having a cord PbB level of 2.88 µmol/litre
    (60 µg/dl) and a PbB level of 3.45 µmol/litre (72 µg/dl) at 2 weeks.
    The effect was seen at 6 weeks and at 3 months but not later
    (chelation therapy was given at 17 days and repeated at 5 months).

         Bauchinger et al. (1977) tested blood taken from 38 children
    selected from a school situated close to a lead plant and in whom the
    PbB level was at least 1.44 µmol/litre (30 µg/dl). There was no
    increase in the frequency of aberrations.

         Grandjean et al. (1983) examined 10 long-term workers at a lead
    storage battery plant. They found normal or slightly increased SCE
    rates in workers with a PbB level of 1.4-3.6 µmol/litre (29-75 µg/dl)
    and a ZPP level of 50-750 µmol/mol haemoglobin. The SCE increase was
    correlated significantly with ZPP level.

         Huang et al. (1988b) studied 21 lead-exposed workers from a
    battery factory. They concluded that the SCE rate in workers with
    long-term exposure to lead increased significantly when the mean PbB

    level was 3.84 µmol/litre (80 µg/dl) or more. There were no
    significant differences in SCE rates among the low- and medium-exposed
    groups or the controls. These authors concluded that the effect of
    chromosome damage caused by lead is not very strong.

         Dalpra et al. (1983) examined blood samples taken from 19
    children living in a contaminated area near a smelter and having PbB
    levels in the range of 1.39-3.02 µmol/litre (29-63 µg/dl). They found
    no effect on SCE frequency.

    8.10  Carcinogenicity

         Much has been written regarding the evidence that lead is
    carcinogenic in humans (Moore & Meredith, 1979; Kazantzis, 1989;
    Goyer, 1992). In several large epidemiological studies no association
    was found which would associate lead with induction of cancer (Cooper
    & Gaffey, 1975; Kang et al., 1980; Cooper, 1981; McMichael & Johnson,
    1982). One major difficulty in many of the studies was the concurrent
    exposure to potential carcinogens such as chromium (Davies, 1984), and
    there has seldom been any attempt to deal with the primary etiological
    agent (smoking) in the development of lung cancers associated with
    lead exposure.

    8.10.1  Occupational exposure and renal cancer

         There are two case reports and very limited other epidemio-
    logical evidence of an association between occupational exposure to
    lead and renal cancer. Confounding variables, including use of tobacco
    and exposure to other carcinogens, were not addressed (Baker et al.,
    1980; Lilis, 1981; Selevan et al., 1985; Cantor et al., 1986).

         The age-standardized mortality ratio for cancer was low in the
    "lead poisoned" and other groups of workers from the Port Pirie study
    by McMichael & Johnson (1982). The authors concluded that lead
    poisoning did not increase the risk of cancer in humans.

    8.10.2  Conclusion

         Although some cohort mortality studies have indicated an
    association between lead exposure and renal disease, there is no
    association between renal cancer and lead in humans.

         A Working Group convened by the International Agency for Research
    on Cancer (IARC) in 1987 concluded that the evidence for the
    carcinogenicity of lead and inorganic lead compounds in humans was
    inadequate.

    8.11  Effects on thyroid function

    8.11.1  Occupational groups

         Robins et al. (1983) found low values for serum thyroxine and
    estimated free thyroxine in 7 of 12 workers having PbB levels above
    2.11 µmol/litre (44 µg/dl). Both measures were correlated with PbB
    level in a cross-sectional study of 47 foundry workers. Serum
    thyrotropin and triiodothyronine levels were within the normal range
    for both study groups.

         No such relationship was found in a study by Refowitz (1984) of a
    one-in-three random sample (N = 58) of male employees at a secondary
    copper smelter. No thyroid abnormalities were observed and there was
    no statistically significant relationship between PbB and serum
    thyroxine or an estimate of free thyroxine in serum.

         Tuppurainen et al. (1988) studied 176 African male workers in
    Kenya and found that the duration of lead exposure correlated
    negatively with serum free thyroxine and serum total thyroxine. The
    correlation was strongest for the workers with the highest exposure
    intensity over time. PbB data were available from periodic medical
    examinations and average values (usually five determinations) were
    used. The mean PbB was 2.73 µmol/litre (56.8 µg/dl), range
    1.01-6.47 µmol/litre (21-135 µg/dl), and there was a mean exposure
    duration of 7.6 years (range 1-20). The authors noted that current PbB
    level, as a point determination, was not associated with total or free
    thyroxine, triiodothyronine or thyrotropin in serum. They proposed
    that long-term, highintensity exposure might be associated with
    depressed thyroid function.

         Gennart et al. (1992b) included assessment of thyroid function as
    part of a study of lead-exposed workers. Data for serum levels of
    triiodothyronine, thyroxine, free thyroxine index and thyroid-
    stimulating hormone were within normal ranges for a group of 98
    workers (mean PbB level, 2.45 µmol/litre or 51 µg/dl) and 85 controls
    (1.00 µmol/litre or 20.9 µg/dl). The lead exposure of this study group
    was considerably lower than in the workers in Kenya, and it was
    suggested that thyroid function changes might not be indicators of
    moderate exposure to lead.

    8.11.2  Effects in children

         Siegel et al. (1989) tested 68 children for thyroid function and
    for PbB and found no statistically significant relationship between
    lead and total or free thyroxine.

    8.12  Immune system

         In a review of the effects of lead on the immune responses of
    experimental animals, Koller (1985) noted that there were few human
    studies.

    8.12.1  Occupational exposure

         There is some evidence that lead workers with PbB levels in the
    range of 1.0-4.10 µmol/litre (21-85 µg/dl) have increased
    susceptibility to infections (colds and influenza) and have a
    significant suppression of secretory IgA levels, a major factor in
    defence against respiratory and gastrointestinal infections (Ewers et
    al., 1982). There are also reports of impaired mitogen responses
    (reflecting T-lymphocyte function) to phytohaemagglutinin (Jaremin,
    1983) and of increased numbers of suppressor T-cells (Cohen et al.,
    1989). However, on the basis of a study of 39 workers exposed to lead
    oxide (mean PbB level of 1.84 µmol/litre or 38.5 µg/dl) and 21 control
    subjects (mean PbB level of 0.57 µmol/litre or 11.8 µg/dl), Kimber et
    al. (1986) concluded that chronic lead exposure in man is not
    associated with the immunological changes that have been observed in
    rodent studies.

         Coscia et al. (1987) found increased B-lymphocyte percentage and
    absolute count in workers currently exposed to lead with PbB levels
    exceeding 2.4 µmol/litre (> 50 µg/dl).

    8.12.2  Children

         A study of 12 pre-school children (Reigart & Graber, 1976) with
    PbB levels > 1.92 µmol/litre (40 µg/dl) did not reveal altered
    immunity in comparison with a control group.

    8.13  Effects on bone

         The pharmacokinetics of the transfer of bone lead to other target
    organs and blood has been discussed by Smith & Hursh (1977), Marcus
    (1985a,b), Silbergeld et al. (1988) and Rabinowitz (1991) (see section
    6.2.2).

         Bone homoeostasis depends upon a complex interaction of its
    various components, i.e. minerals, cells and the extracellular matrix
    composed of collagenous and non-collagenous proteins (Marks & Popoff,
    1988; Sauk & Somerman, 1991). Early work on the effects of lead on
    calcium homoeostasis and calcium-mediated function was reviewed by
    Pounds (1984). More recent work has been summarized by Pounds et al.
    (1991), particularly the effects of lead on hormone action,
    competition with calcium binding on calcium messenger systems, and the
    impaired synthesis of collagen or sialoproteins. Additional evidence

    for adverse effects of lead on the proteins in the mineral
    compartment, thus affecting bone homoeostasis, has been reported by
    Sauk & Somerman (1991).

    8.14  Biomarkers for lead effects

         The understanding and application of biomarkers for the
    assessment of effect is becoming more widespread (IPCS, 1994).

         It is desirable to identify biomarkers of lead exposure and
    effect to provide easily measurable parameters that will facilitate
    the risk assessment process. At present PbB levels are frequently
    measured to assess both exposure and effect. Alternative biomarkers
    for lead which may be easily measured are of biochemical effects,
    particularly in the haem biosynthetic pathway.

         However, the relationship between these effects and neurological
    impairment caused by lead has not been established. Measurement of
    these effects has also been used to provide biomarkers for exposure.

         Anaemia has been associated with high levels of lead exposures in
    the occupational setting. However, clinical tests for haemoglobin are
    rather non-specific indicators of lead toxicity (Bernard & Becker,
    1988). The steps within the haem biosynthetic pathway which have been
    used to measure effect are: 1) inhibition of delta-aminolaevulinic
    acid dehydratase (ALAD); 2) urinary excretion of delta-aminolaevulinic
    acid (ALAU); (3) the accumulation of zinc protoporphyrin (ZPP) in
    erythrocytes arising from the inhibition of the enzyme ferrochelatase
    or the iron transport system.

         There are several methodological problems related to the
    measurement of ALAD, and the benefit of measuring enzyme activity in
    blood in lieu of PbB determinations is not apparent (Mushak, 1989).
    Early work (Roels et al., 1976) indicated enzyme inhibition at PbB
    levels of 0.24 µmol/litre (> 5 µg/dl). However, the data at levels of
    0.72 µmol/litre (< 15 µg/dl) appear scattered (Hernberg & Nikkanen,
    1970; Granick et al., 1973).

         Levels of ALA in urine have been used as an effect indicator of
    lead exposure (Meredith et al., 1978). ALAU levels were found to
    correlate with PbB levels as low as 0.864 µmol/litre (18 µg/dl), the
    correlation becoming much stronger at PbB levels above 1.92 µmol/litre
    (40 µg/dl) (Selander & Cramer, 1970). Given the present concern over
    effects in children at PbB levels well below 1.92 µmol/litre
    (40 µg/dl), the assay of ALAU concentrations may not be sensitive
    enough to be of any value.

         The effects of lead on porphyrin levels, both coproporphyrin in
    urine and zinc protoporphyrin (ZPP) in blood, have been investigated
    as possible biomarkers for lead. Measurement of urinary coproporphyrin

    levels does not appear to be sensitive enough to be useful in
    assessing exposure to lead at present environmental levels. Excretion
    levels do not rise significantly until PbB levels exceed
    1.92 µmol/litre (40 µg/dl) (Meredith et al., 1978). Other studies
    indicate that this assay may not be sensitive enough for biological
    monitoring at PbB levels between 0.96 and 1.20 µmol/litre (20 and
    25 µg/dl). At present it would appear that ZPP has limited usefulness
    as a biological indicator of exposure at exposures leading to PbB
    levels lower than 1.20 µmol/litre (25 µg/dl) (Roels et al., 1976;
    Piomelli et al., 1982; Hammond et al., 1985; Marcus & Schwartz, 1987).
    The results may also be confounded by concurrent iron deficiency which
    will alter the levels of ZPP.

         Vitamin D metabolism involves cytochrome P-450-dependent enzymes
    and thus may be affected by lead exposure. Measurement of serum levels
    of 1,25-dihydroxy-vitamin-D has been proposed as a sensitive
    biological monitor of lead exposure in children (Rosen et al., 1980;
    Mahaffey et al., 1982; Koo et al., 1991). However, it should be noted
    that dietary intakes of calcium and phosphorus, as well as circulating
    levels of parathyroid hormone will regulate the production and
    circulating concentrations of this vitamin D metabolite, thus making
    correlations with PbB levels difficult without adequate information on
    the nutritional status of the population under study (Rosen & Chisney,
    1983). In populations showing adequate nutritional status, lead at
    exposures leading to PbBs levels lower than 0.96 µmol/litre (20 µg/dl)
    does not appear to have a demonstrable effect on circulating levels of
    1,25-dihydroxy-vitamin-D.

    9.  EVALUATION OF HUMAN HEALTH RISKS

    9.1  Exposure assessment

         Lead is a ubiquitous element detected in all environmental media.
    However, natural sources contribute only a small fraction of the
    amounts of lead found in air, food, water and dust. The majority of
    lead in these media arises from automobile and industrial emissions
    and from the use of lead-containing solder and paints. Adults and
    older children receive the largest proportion of lead intake from
    foods, whereas dust, soil and food all make significant contributions
    to the total lead intake of young children. The major contributions to
    lead in soil and outdoor dust are from the combustion of fossil fuels
    (principally leaded petrol), stationary sources such as smelters, and
    peeling and flaking of lead-based paint.

    9.1.1  General population exposure

         In the absence of specific stationary sources of lead,
    concentrations in ambient air are directly related to density of
    traffic and whether lead is still utilized as an additive in petrol.
    Reduction or elimination of lead in petrol in those countries which
    have instituted regulations has resulted in a decline by as much as
    eight-fold in ambient air concentrations of lead.

         Levels of lead in indoor air are affected by the presence of
    cigarette smoke and dust from lead-painted surfaces. Without such
    sources, air lead levels indoors are about 60% of those in outdoor
    air.

         For most adults, the total daily exposure to lead is via food,
    water and air. For infants aged up to 5 months, formula or breast milk
    and water are the main sources of lead. In children, an additional
    source of exposure is dust and soils. Absorption is dependent on the
    chemical form of lead, type of soil and particle size
    (bioavailability). Lead intake may be augmented from unusual sources
    such as folk remedies, cosmetics and hobby activity. Community
    contamination and workplace practices may contribute to lead exposure.

         Food (including drinking-water and beverages) is the major source
    of lead exposure for the general population. Infants and children may
    receive an added lead burden from soil and dust. The most significant
    foodstuffs will vary from country to country. In areas still utilizing
    lead-soldered cans, levels of lead are substantially higher. Depending
    upon lifestyles, there may be significant oral intake of lead from
    some alcoholic beverages and due to the leaching of lead from low
    temperature-fired ceramic containers.

         Most drinking-water supplies contain lead levels lower than
    5 µg/litre when they leave the treatment plant. However, where the
    water is known to be plumbo-solvent, up to 40% of the samples may
    exceed 100 µg/litre in homes where lead solder, lead pipes or brass
    fixtures have been used.

         Absorption of lead from the lung is a function of particle size
    and pulmonary deposition pattern. Small particles (< 0.5 µm in
    diameter) characteristic of ambient air will be deposited deeply in
    the lungs with absorption rates of 90%. Larger particles, such as
    those that may be encountered in occupational settings, exhibit high
    deposition rates in the upper airway. Absorption of such particles
    will be a function of both dissolution in the lung and particle
    clearance to the gastrointestinal tract.

         Human dermal absorption of inorganic lead through unabraded skin
    is of limited significance.

    9.1.2  Occupational exposures

         In addition to exposure to lead in ambient air, water and food,
    some workers may be exposed to airborne lead and dust within the
    workplace. Actual levels will vary according to the engineering design
    of the process equipment and the industrial hygiene practices
    utilized.

    9.2  Critical issues related to exposure evaluation

         In view of the heterogeneity of responses to lead within human
    populations, the complex interrelationships between exposure to lead
    and a biological indicator for internal dose require consideration of
    several key issues in order to assess human exposures.

    9.2.1  Sampling and analytical concerns

         Reliable comparisons of reported levels of exposure and/or dose
    can only be made where authors have described the analytical and
    sampling procedures in sufficient detail to allow the reader to
    assess, for example, the integrity of the sampling procedures as well
    as the specificity, precision and accuracy of all analytical methods.
    Problems related to sampling cord blood also need to be considered.

         Most studies have utilized analytical procedures of high quality.
    However, consideration must be given to blood collection procedures
    (finger stick versus venepuncture) when comparing results.

    9.2.2  Data presentation

         Inter-study comparisons of lead exposures are complicated by the
    variety of methods used to present results, including median values as
    well as geometric and arithmetic means. Some authors have used
    log-transformed data.

         In assessing exposure from data on teeth, one must know which
    tooth and which compartment(s) within the tooth were sampled. In
    addition, if data are to be compared between studies, authors must
    state explicitly that all teeth analysed were without caries and were
    shed spontaneously.

         The exposure index lifetime average blood lead (PbB) level has
    assisted in the assessment of cumulative exposures from serial PbB
    data. It should not be interpreted as being equivalent to a single PbB
    determination at a single point in life.

         For some data/analytical purposes, age-specific PbB levels may be
    more appropriate than a lifetime average.

    9.3  Relationship between exposure and dose

         The most widely used surrogate for the absorbed dose is whole PbB
    concentration.

         The relationship between PbB level and lead intake is curvilinear
    over a wide range of PbB values. On the basis of a single study of 17
    infants, the relationship between PbB level and lead intake from food
    has been determined to be 0.0077 µmol lead/litre (0.16 µg/dl) per µg
    lead intake per day for a median PbB level of approximately
    0.48 µmol/litre (10 µg/dl).

         Most studies of the relationship between PbB level and lead
    exposure apply to a single environmental source, i.e. air, food, water
    or soil/dust. A summary of the relationship between PbB level and lead
    intake from individual media is given in Table 23.

    9.4  Surrogate measures of dose

    9.4.1  Blood

         Whole PbB values are widely used as a measure of absorbed dose.
    However, it is believed that plasma lead concentrations may better
    reflect the "active" fraction of lead in blood and define the
    relationship between PbB and tissue or organ accumulation (and

        Table 23.  Representative relationships of blood lead median level to intake of
               lead for the general populationa
                                                                                   
                                          Population
                                                                                   
    Median         Children                           Adults
                                                                                   

    Airb           0.09 µmol Pb/litre per µg          0.079 µmol Pb/litre per µg
                   Pb/m3 (1.92 µg Pb/dl)              Pb/m3c (1.64 µg Pb/dl)

    Water                                             0.003 µmol Pb/litre per µg
                                                      Pb/litre (0.06 µg Pb/dl)

    Food           0.01 µmol Pb/litre per µg          0.002-0.003 µmol Pb/litre per
                   Pb/day (0.16 µg Pb/dl)             µg Pb/day (0.04-0.06 µg Pb/dl)

    Dustb          0.09 µmol Pb/litre per 1000 µg
                   Pb/g (1.8 µg Pb/dl)

    Soilb          0.11 µmol Pb/litre per 1000 µg
                   Pb/g (2.2 µg Pb/dl)
                                                                                   

    a  These data are provided for illustrative purposes only recognizing that the
       relationships are curvilinear in nature and are broad guidelines which will
       not apply at lower or higher levels of exposure.

    b  A value between 0.144 to 0.24 µmol Pb/litre or 3-5 µg Pb/dl per µg/m3
       is obtained when one considers indirect contribution through deposition on
       soil/dust.

    c  The air to blood lead relationship in occupational settings is best described
       by a curvilinear relationship having slopes between 0.02 and 0.08 µg/m3
       air. The slope is variable but lower than that found for humans in the general
       environment, which is between 1.6 and 1.9 µg/m3.
    

    effect), although there is little experimental data (because of
    analytical limitations). PbB is distributed between plasma and
    erythrocytes, with less than 5% being in the plasma. Most of the lead
    is bound to haemoglobin.

         Venous and capillary blood levels are generally equivalent,
    provided that the sampling technique is adequate.

    9.4.2  Urine

         Urinary measurements of lead concentration are of limited value,
    although they are used occasionally as a screening test for
    occupational population groups.

    9.4.3  Bone

         Bone lead may be measured by non-invasive X-ray fluorescence, but
    this technique is limited in sensitivity at present.

    9.4.4  Tooth

         Shed deciduous teeth have been used to provide an index of
    exposure in early childhood. Interpretation of the analytical data is
    dependent on the type of tooth and the part of the tooth (whole tooth,
    dentine or circumpulpar dentine) analysed.

    9.4.5  Hair

         Hair is not useful for measurement of lead exposure.

    9.5  Biochemical effects of lead

    9.5.1  Haem synthesis

         Evaluation of the quality of analytical data is an important
    aspect in considering reports describing effects attributed to lead.
    It should be noted that much of the data presented in this area has
    not been as vigorously scrutinized as, for example, psychometric study
    data.

         An increase in erythrocyte protoporphyrin (EP) in children occurs
    between PbB levels of 0.72 and 1.2 µmol/litre (15-25 µg/dl). Increases
    in EP can be detected in men when the PbB level is above
    1.20-1.44 µmol/litre (25-30 µg/dl), and in women when it is above
    0.96-1.44 µmol/litre (20-30 µg/dl). It should be noted that the effect
    of lead on haem is confounded by low iron status.

    9.5.1.1  Urinary coproporphyrin

         The coproporphyrin concentration in urine increases significantly
    with PbB levels in excess of 1.92 µmol/litre (40 µg/dl).

    9.5.1.2  Urinary aminolaevulinic acid in children

         In children 1-5 years old, there is a linear relationship with
    PbB in the range 1.2-3.6 µmol/litre (25-75 µg/dl). Data for children
    with PbB levels of 0.24-1.92 µmol/litre (5-40 µg/dl) show essentially

    no correlation with urinary aminolaevulinic acid (ALA) excretion.
    Elevation of urinary ALA level is evident at a PbB level of about
    1.68 µmol/litre (35 µg/dl).

    9.5.1.3  Urinary aminolaevulinic acid in adults

         Urinary excretion of ALA increases in men at PbB levels above
    2.16 µmol/litre (45 µg/dl) and in women above 1.68 µmol/litre
    (35 µg/dl).

    9.5.1.4  delta-Aminolaevulinic acid dehydratase

         There was a negative exponential relationship between PbB level
    and delta-aminolaevulinic acid dehydratase (ALAD) activity in a
    population of 10- to 13-year-old children with PbB levels in the range
    of 0.19-1.97 µmol/litre (4.7-41 µg/dl). An effect was seen at a PbB
    level of approximately 0.48 µmol/litre (10 µg/dl). There is an
    apparent lack of a clearly defined threshold for lead inhibition of
    ALAD in different age groups.

    9.5.2  Vitamin D metabolism

         In the presence of adequate nutritional status, PbB levels below
    0.96 µmol/litre (20 µg/dl) appear to have no demonstrable effect on
    circulating concentrations of 1,25-dihydroxy-vitamin-D. A PbB level
    above 0.96 µmol/litre (20 µg/dl) is associated with a decrease in the
    serum level of 1,25-dihydroxy-vitamin-D.

    9.5.3  Dihydrobiopterin reductase

         Inhibition of dihydrobiopterin reductase has been shown in humans
    where the mean PbB level is as low as 0.48 µmol/litre (10 µg/dl).

         A summary of some biochemical effects of lead is presented in
    Table 24.

    9.5.4  Haemopoietic system

    9.5.4.1  Anaemia in adults

         The estimated PbB associated with a decrease in haemoglobin
    concentration is 2.40 µmol/litre (50 µg/dl).

    9.5.4.2  Anaemia in children

         Decreased haemoglobin levels in children occur at a PbB level of
    approximately 1.92 µmol/litre (40 µg/dl). Anaemia, defined as a
    haematocrit below 35%, is not found at a PbB level of less than
    0.92 µmol/litre (20 µg/dl). The risk of having a haematocrit value
    below 35% for a 1-year-old child is 2% at a PbB level of

    Table 24.  Biochemical effects of lead
                                                                       
    Parameter                    Blood lead level above which the
                                 biochemical effect is demonstrable with
                                 current techniques

                                      µmol/litre             µg/dl
                                                                       

    Protoporphyrin levels             0.96-1.44              20-30

    Coproporphyrin levels                1.92                  40

    ALA urine levels                     1.68                  35

    ALAD activity                        0.48                  10

    1,25-dihydroxy-vitamin-D             0.96                  20

    Dihydrobiopterin reductase           0.48                  10
                                                                       


    0.96-1.87 µmol/litre (20-39 µg/dl); the contribution of iron
    deficiency may account for a substantial proportion of this 2%.
    Induction of anaemia is demonstrable at 1.92 µmol/litre (40 µg/dl).

    9.5.4.3  Erythrocyte pyrimidine-5'-nucleotidase

         A reduction of 20% or more in erythrocyte pyrimidine
    -5'-nucleotidase activity is associated with PbB concentrations above
    0.48 µmol/litre (10 µg/dl).

         Effects of lead are demonstrable on a number of enzyme systems
    and biochemical parameters. The PbB levels, above which effects are
    demonstrable with current techniques for the parameters which may have
    clinical significance, are all greater than 0.96 µmol/litre
    (20 µg/dl). Some clinically insignificant effects on enzymes are
    demonstrable at lower levels of PbB.

    9.6  Nervous system

    9.6.1  Adults

    9.6.1.1  Central nervous system

         With acute lead exposure resulting in a PbB level in excess of
    3.84 µmol/litre (80 µg/dl), severe encephalopathy and/or coma may
    occur. Central nervous system (CNS) symptoms are found in lead-exposed
    adults when there is a history of several years of exposure to lead at

    PbB levels that may not have exceeded 3.36 µmol/litre (70 µg/dl) and a
    PbB level at the time of clinical assessment of at least
    1.92 µmol/litre (40 µg/dl).

         Impaired neurobehavioural test performance has been found in
    lead-exposed workers. Changes in critical flicker fusion test have
    been detected at a PbB level of about 2.4 µmol/litre (50 µg/dl).
    Sensory motor function is generally more sensitive than cognitive
    end-points in many neurobehavioural evaluations, the lowest-observed-
    effect level being at about 1.92 µmol/litre (40 µg/dl).

         It appears also that neuroelectrophysiological tests are
    sensitive indicators of the CNS effects of lead. Reductions in
    latencies of sensory evoked potentials and auditory event-related
    potentials have been found in workers with average PbB levels of
    approximately 1.92 µmol/litre (40 µg/dl).

    9.6.1.2  Peripheral nervous system

         Numerous studies have measured the conduction velocity of
    electrically stimulated sensory and motor nerves in workers exposed to
    lead. These nerve conduction velocity (NCV) studies have yielded
    somewhat mixed results, with many showing a decrease in NCV in
    relation to lead exposure (indexed as PbB level) and a few showing no
    effect or occasionally even an increase in NCV associated with lead
    exposure. Differences in the nerves evaluated, methodologies,
    characterization of lead exposure, and control of confounding
    variables underlie some of the variability in results across studies.
    A statistical meta-analysis of 32 NCV studies has indicated that NCV
    is significantly reduced in lead-exposed workers compared to reference
    subjects, but that the median motor nerve shows more reliable effects
    of lead than other nerves. This collective view of the evidence is
    supported by key studies that provide compelling evidence of a causal
    relationship between lead exposure and reductions in NCV, extending to
    PbB levels as low as 1.44 µmol/litre (30 µg/dl). These effects may be
    reversible depending on the duration and level of exposure.

    9.6.1.3  Autonomic nervous system

         Two reports examining the electrocardiographic RR interval
    variability during deep breathing and the component CV of respiratory
    sinus arrythmia demonstrated dysfunctions at an average PbB level of
    1.68 µmol/litre (35 µg/dl). These results suggest autonomic nervous
    system dysfunction, particularly the parasympathetic nervous system.

    9.6.2  Children

         Prospective and cross-sectional studies of children have
    demonstrated associations of lead exposure, measured by various

    indices, and intellectual performance. The association has been noted
    across a wide range of exposure levels and in a variety of populations
    before factors other than lead have been accounted for.

         A key question is whether this statistical association is
    directly attributable to the causal influence of lead on child IQ. It
    is important to consider alternative contributory explanations, i.e.
    random chance, unexplained confounding factors, reverse causality and
    selection bias.

         It is a matter of debate and conjecture as to the extent to which
    these four issues should inhibit claims of a causal relationship in
    the epidemiological studies. The essential problem is that
    observational epidemiology cannot provide definitive evidence of
    causality when the key statistical association is small, the temporal
    relationship is unclear and major confounders are present. Animal
    studies provide qualitative support for the claim of a causal role for
    lead in affecting neuropsychological performance, but provide limited
    assistance in establishing quantitative dose-effect relationships.

    9.6.2.1  Type of effect

         The clearest and most consistent associations have been found
    with global measures, such as IQ, where the largest body of evidence
    is available. Efforts to delineate the neuropsychological foundations
    of this association with a wide variety of tests of specific
    neuropsychological domains have not so far been successful.

    9.6.2.2  Magnitude

         Based on the evidence from cross-sectional and prospective
    studies of populations with PbB levels generally below 1.2 µmol/litre
    (25 µg/dl), the size of the apparent IQ effect (at ages 4 and above)
    is a deficit of between 0 and 5 points (on a scale of 100 with a
    standard deviation of 15) for each 0.48 µmol/litre (10 µg/dl)
    increment in PbB level, with a likely apparent effect size of between
    1 and 3 points. At PbB levels above 1.2 µmol/litre (25 µg/dl), the
    relationship between PbB level and IQ may differ. Estimates of effect
    size are group averages and only apply to the individual child in a
    probabilistic manner.

         Existing epidemiological studies do not provide definitive
    evidence of a threshold. Below the PbB range of 0.48-0.72 µmol/litre
    (10-15 µg/dl), the effect of confounding variables and limits in the
    precision of analytical and psychometric measurements increases the
    uncertainty attached to any estimate of effect. However, there is some
    evidence of an association below this range.

    9.6.2.3  Reversibility/persistence

         Whilst the IPCS Task Group could not unequivocally state that
    effects of early childhood exposure are persistent beyond childhood,
    because the current data are too meagre, it was held that
    neurobehavioural effects detected at age seven or later usually
    persist. Measures in later childhood tend to be more predictive of
    subsequent performance than those made earlier. It is more likely than
    not that effects seen during school years are to some degree
    irreversible. This has also been observed in later follow-up studies
    conducted in other non-lead topics of child development research. One
    of the difficulties is that there are too few studies concerning
    long-term outcome in children with high early exposures and where the
    sources of exposure are subsequently removed.

         Virtually no useful data are available on the effects on IQ of
    removing children from a "high" exposure environment to one of "low"
    exposure or on reduction of body lead burden in children. This is not
    to say that exposure should not be reduced when possible.

    9.6.2.4  Age-specific sensitivity

         From prospective studies it is not possible to determine an age
    of critical sensitivity. This reflects the findings that serial PbB
    measures taken at the age of 2 years and later are positively
    correlated with, the individual rankings remaining approximately
    constant, and this limits the ability to identify sensitive periods of
    exposure.

    9.6.2.5  Interactions/subgroups

         The evidence is inconclusive on whether apparent effects are more
    or less marked in different gender or socioeconomic status (SES)
    subgroups. However, where there are suggestions of SES-related
    differences, the apparent effects tend to be more marked in the lower
    SES subgroups.

    9.6.3  Animal studies

         Experimental animal studies of CNS effects provide support for
    the associations between PbB levels and neurobehavioural deficits
    described in human epidemiological studies of lead. There is
    supportive evidence both in terms of demonstrating causal
    relationships and in the levels of PbB at which such effects are
    observed, namely 0.528-0.72 µmol/litre (11-15 µg/dl). Moreover, they
    provide qualitative parallels in the nature of the effects described,
    as these effects include changes in learning and memory functions.
    Experimental animal studies indicate that these CNS effects may depend
    upon task complexity and can persist long beyond the termination of
    lead exposure. These studies also provide information possibly

    relevant to understanding mechanisms of effect. In addition, the
    experimental animal studies provide such evidence in the absence of
    the confounding factors and co-variates, such as parental IQ,
    socioeconomic status, and quality of the home environment, that are
    problematic to human epidemiological endeavours, and in the absence of
    nutritional deficiencies that may arise in human populations.

    9.7  Renal system

         Renal function impairment was not associated with a PbB level
    below 3.0 µmol/litre (62 µg/dl) when measured by blood urea nitrogen
    and serum creatinine levels in lead workers. However, renal tubular
    effects were detected in workers with a PbB level below 3.0 µmol/litre
    when measured by more sensitive indicators such as NAG.

         Most studies of the general population attempting to relate renal
    function impairment to PbB concentration have not demonstrated an
    effect with PbB levels below 1.8 µmol/litre (37.3 µg/dl). More
    sensitive indicators of renal function may indicate a renal effect of
    lead below this level.

    9.8  Liver

         Over-exposure to lead may inhibit drug metabolism in the liver.

    9.9  Reproduction

    9.9.1  Female

         Studies on the risk of spontaneous abortion and reduced birth
    weight associated with maternal PbB levels below 1.44 µmol/litre
    (30 µg/dl) have yielded mixed results. Recent epidemiological studies
    have shown exposure-related perturbations in the length of gestation,
    significantly greater risks being associated with PbB levels of
    0.72 µmol/litre (15 µg/dl) or more.

    9.9.2  Male

         PbB concentrations above 1.92 µmol/litre (40 µg/dl) have been
    shown to affect sperm morphology and function. At present, the
    reproductive consequences of these changes are unknown.

    9.10 Blood pressure

         A quantitative assessment of the collective evidence from all the
    observational studies in adults is made difficult by the fact that
    studies have adopted different policies regarding adjustment for
    potential confounding factors (e.g., alcohol consumption). In
    addition, quantitative findings from the two largest studies (BRHS and
    NHANES II) have depended on whether adjustment was made for
    geographical variations in blood pressure and blood lead.

         The limited size of most observational studies has inevitably
    meant that they could not consistently demonstrate a statistically
    significant relationship. However, an overview of all the studies
    shows that evidence is consistent with the centre-adjusted analysis of
    the two main studies, i.e. there are very weak but statistically
    significant associations between PbB level and both systolic and
    diastolic blood pressures. The likely order of magnitude is that for
    any two-fold increase in PbB level (e.g., from 0.8 to 1.6 µmol per
    litre) there is a mean 1 mmHg increase in systolic blood pressure. The
    association with diastolic blood pressure is of a similar magnitude.

         Animal studies have provided plausible mechanisms for an effect
    of lead on blood pressure. However, from such a small magnitude of
    statistical associations in the presence of important confounders, one
    cannot infer that low-level lead exposure is causally related to an
    increase in blood pressure.

         The two population studies relating PbB to cardiovascular disease
    events show no statistically significant associations. Hence, there is
    no clear evidence to suggest that lead has an impact of public health
    importance as regards hypertension or risk of cardiovascular disease.

    9.11  Carcinogenicity

         Renal tumours occur in rats and mice administered high doses of
    lead. However, the evidence for the carcinogenicity of lead and
    inorganic lead compounds in humans is inadequate.

    9.12  Immune system

         There is no strong evidence in humans of an effect of lead on the
    immune system.

    10.  RECOMMENDATIONS FOR PROTECTION OF HUMAN HEALTH

    10.1  Public health measures

         Public health measures should be directed towards reduction and
    prevention of exposure to lead by reducing the use of lead and lead
    compounds and by minimizing lead-containing emissions that result in
    human exposures. This can be achieved by:

    a)   phasing-out any remaining uses of lead additives in motor fuels;

    b)   further reducing the use of lead-based paints, with the objective
         of eliminating such paints;

    c)   development and application of methods for the safe and
         economical remediation of lead-painted homes and lead
         contaminated soil;

    d)   elimination of the use of lead in food containers (e.g., in the
         seams of cans);

    e)   dissemination of information to assist with identification of
         glazed food containers which may leach lead into food placed,
         cooked or stored in the container;

    f)   eliminating any remaining agricultural uses of lead or lead
         compounds (e.g., lead arsenate as an insecticide);

    g)   identifying and reducing, or preferably, eliminating lead found
         as a contaminant or ingredient of folk remedies and cosmetics;

    h)   the use of materials and engineering practices to minimize
         plumbosolvency in water treatment and water distribution systems;

    i)   systematic examination of processes in which lead is used or
         recycled in order to identify and reduce lead exposure by means
         of improved engineering design, of operators, by-standers and the
         environment. Opportunities for technology transfer should be used
         whenever possible.

    10.2  Public health programmes

         Programmes should be developed:

    a)   to enhance data collection and to make available to the public
         information on the lead content of foodstuffs;

    b)   to facilitate identification of populations at high risk of
         exposure to lead on the basis of monitoring data for lead in
         food, air, water and soil;

    c)   that incorporate improved procedures for health risk assessment
         of population groups at risk of exposure to lead;

    d)   that promote understanding and awareness concerning the effects
         on human health associated with exposure to lead, while
         recognizing cultural sensitivities;

    e)   that place emphasis on adequate nutrition, health care and
         attention to socioeconomic conditions which may exacerbate the
         effects of lead present in the environment.

    10.3  Screening, monitoring and assessment procedures

         Methods for evaluating the effects and associated risks of
    exposure to lead require both improvement and further development or
    research. In the short term the following measures are needed.

    a)   Screening

    i)   Blood lead measurements should be recognized as the biomarker of
         choice for screening for previous exposure of children to lead.

    ii)  The sensitivity of the developing nervous system to the
         potentially harmful effects of lead is such that other
         biochemical measurements (e.g., erythrocyte protoporphyrin) are
         not sufficiently sensitive for assessment of infants and
         children.

    b)   Monitoring

    i)   More sensitive analytical methods should be developed for the
         reliable measurement of blood lead levels below 0.72 µmol/litre
         (15 µg/dl) to acceptable standards of precision and accuracy.

    ii)  There is a need for international analytical quality assurance
         programmes utilizing reference lead-containing materials.

    iii) All publications containing blood lead measurement data should
         provide adequate data on current quality assurance and quality
         control.

    iv)  Data comparisons are made more difficult by differences in units
         and statistical techniques for data handling. Investigators
         should be encouraged to adopt internationally agreed practices
         (e.g., IUPAC units).

    c)   Assessment

    i)   Validated biomarkers are needed to define the relationship
         between measures of environmental (external) exposure and
         specific effects (biochemical or functional).

    ii)  Biomarkers indicative of deficits in cognitive performance are
         needed to facilitate assessment of risk.

    iii) Improved methods are needed for the definition of outcome effects
         (especially IQ and neurobehavioural deficits) attributable to
         lead at blood lead concentrations of about
         0.48 µmol/litre (10 µg/dl) or less.

    iv)  Further data are required to determine whether outcome effects on
         the nervous system attributed to lead are reversible or
         permanent.

    v)   Biomarkers for the renal effects of lead are needed to link renal
         damage with lead exposure and thereby improve assessment of risk
         of renal damage.

    11.  FURTHER RESEARCH

    a)   Research is required to improve the process for the assessment of
         risk associated with exposure to lead. Specifically this should
         cover:

         i)     definition of the health significance of biochemical
                changes associated with exposure to lead, with particular
                attention to alterations associated with blood lead
                concentrations of about 0.72 µmol/litre (15 µg/dl) or
                less;

         ii)    work to define the bioavailability of lead from different
                sources and to establish the relationship between exposure
                (sources and speciation) and body burden;

         iii)   definition of the influence of host-related factors
                (particularly nutrition) affecting absorption and
                distribution of lead;

         iv)    intensification of kinetic studies of lead to provide an
                improved data base for extrapolation between species;

         v)     elucidation of mechanisms of accumulation and mobilization
                of lead from bone with particular attention to the
                influence of pregnancy and ageing on kinetics;

         vi)    investigation of the pharmacokinetics of lead in the
                pregnant female in relation to transfer of lead to the
                developing embryo and fetus and factors that mitigate such
                transfer;

         vii)   determination of the effects of pre- and post-natal
                exposure to lead;

         viii)  improved definition of paternally mediated effects of lead
                exposure on the reproductive process and outcomes.

    b)   More general research needs include the following:

         i)     rationalization of neurobehavioural tests used to assess
                performance in children and animals in order to permit
                comparisons of measurements and test data that use similar
                biological mechanisms;

         ii)    follow-up research (where possible) of historical cohorts
                of lead-poisoned adults to examine the sequelae;

         iii)   evaluation of intervention measures, such as lead
                abatement from home and soil, and chelation on reduction
                of blood lead levels and outcome effects on the central
                nervous system.

    12.  PREVIOUS EVALUATIONS BY INTERNATIONAL BODIES

         The carcinogenic potential of lead and lead compounds was last
    evaluated by the International Agency for Research on Cancer in 1987
    (IARC, 1987b). There was inadequate evidence for the carcinogenicity
    of lead and inorganic lead compounds in humans, but sufficient
    evidence was available to show that specified inorganic lead compounds
    were carcinogenic in experimental animals. The overall evaluation
    placed lead and inorganic lead compounds in Group 2B, i.e. possibly
    carcinogenic to humans.

         To protect workers from the adverse effects of lead on haem
    synthesis and on the peripheral and central nervous system, a
    health-based biological exposure limit of 1.92 µmol/litre (40 µg/dl)
    was recommended by a WHO Study Group (WHO, 1980). It was further
    recommended that blood lead (PbB) levels in women within the
    reproductive age range should not exceed 1.44 µmol/litre (30 µg/dl).
    Depending upon the background levels of PbB in the worker population,
    air lead levels should not exceed 30-60 µg/m3 (WHO, 1980).

         A drinking-water guideline value of 0.050 mg/litre was developed
    in 1984 (WHO, 1984). This guideline value has been revised recently to
    0.01 mg/litre (WHO, 1993).

         Lead was evaluated by a WHO Working Group developing air quality
    guidelines for Europe (WHO, 1987). Based on the assumption that PbB
    levels in 98% of the population would be maintained at levels below
    0.96 µmol/litre (20 µg/dl), a guideline value in the range of
    0.5-1.0 µg/m3 (long-term average, such as annual mean) was
    recommended.

         At the 41st meeting of the Joint FAO/WHO Expert Committee on Food
    Additives and Food Contaminants, a Provisional Tolerable Weekly Intake
    (PTWI) of 25 µg/kg body weight was recommended (FAO/WHO, 1993). This
    level refers to lead from all sources and was set to protect all
    humans, including infants and children. It was based on a model
    indicating daily intakes of lead between 3-4 µg/kg body weight by
    infants and children and is not associated with an increase in PbB
    concentrations.

         Regulatory standards established by national bodies in several
    countries and the European Economic Community are summarized in the
    legal file of the International Register of Potentially Toxic
    Chemicals (IRPTC, 1987).

    REFERENCES

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    RESUME

         La présente monographie est consacrée aux risques pour la santé
    humaine découlant de l'exposition au plomb et aux dérivés minéraux du
    plomb. On y insiste sur les données dont on a eu connaissance depuis
    la publication du No 3 de la Série des Critčres de l'environnement:
    Plomb (OMS, 1977). Quant aux effets du plomb au niveau de
    l'environnement, ils sont étudiés dans le No 85 de cette série: Lead -
    Environmental Aspects (Le plomb dans l'environnement) (WHO, 1989, en
    anglais seulement).

    1.  Identité, propriétés physiques et chimiques et méthodes d'analyse

         Le plomb est un métal mou, gris argent, qui fond ŕ 327,5°C. Il
    est trčs résistant ŕ la corrosion mais se dissout ŕ chaud dans l'acide
    nitrique et l'acide sulfurique. La valence habituelle du plomb dans
    ses dérivés minéraux est de +2. La solubilité dans l'eau de ses
    composés est variable, le sulfure et les oxydes étant peu solubles et
    le nitrate, le chlorate et le chlorure assez solubles ŕ chaud. Le
    plomb forme également des sels avec des acides organiques tels que
    l'acide lactique et l'acide acétique ainsi que des composés organiques
    stables tels que le plomb-tétraéthyle et le plomb-tétraméthyle.

         Les méthodes d'analyse les plus couramment utilisées pour doser
    le plomb ŕ faible concentration dans les produits biologiques et dans
    l'environnement sont la spectrométrie d'absorption avec atomisation en
    flamme, en four ŕ cellules de graphite et en plasma induit par haute
    fréquence (ICP/PIHF), ainsi que la coulométrie avec redissolution
    anodique. Selon le traitement préalable subit par l'échantillon et les
    techniques d'extraction et l'instrumentation utilisées, la limite de
    détection peut ętre de 0,12 µmoles de plomb par litre de sang
    (2,49 µg/dl). Toutefois, on n'obtient de résultats fiables qu'en se
    conformant ŕ une marche ŕ suivre bien codifiée pour réduire au minimum
    le risque de contamination au cours du prélčvement des échantillons,
    de leur conservation, de leur traitement et de leur analyse.

    2.  Sources d'exposition humaine

         La proportion du plomb dans la croűte terrestre est d'environ
    20 mg/kg. Le plomb présent dans l'environnement peut ętre d'origine
    naturelle ou artificielle. Le plomb naturellement présent dans
    l'atmosphčre provient, entre autres, de l'altération des roches et des
    éruptions volcaniques, et on l'estime ŕ 19 000 tonnes/an, contre
    126 000 tonnes/an provenant des mines et des fonderies, la
    consommation annuelle totale étant supérieure ŕ 3 millions de tonnes.

         On a trouvé des concentrations de plomb atmosphérique de
    50 pg/m3 dans des zones reculées. La concentration de fond du plomb
    dans le sol varie de 10 ŕ 70 mg/kg et on a relevé ŕ proximité de
    routes une teneur moyenne de 138 mg/kg. Actuellement la teneur de

    l'eau en plomb dépasse rarement quelques microgrammes/litre; la
    concentration naturelle du plomb dans les eaux de surface est estimée
    ŕ 0,02 µg/litre.

         Le plomb et ses dérivés peuvent pénétrer dans l'environnement par
    une porte d'entrée quelconque lors de l'extraction miničre, des
    opérations de fonte et du traitement, de l'utilisation, du recyclage
    ou du rejet du plomb. Le plomb est principalement utilisé dans les
    batteries, les câbles, les pigments, les additifs pour essence, la
    brasure et les aciers. Le plomb et ses dérivés entrent également dans
    la composition de la brasure que l'on utilise pour souder les
    conduites d'eau et les boites de conserve alimentaire; il sert
    également ŕ la confection de certains remčdes traditionnels, des
    capsules de bouteilles contenant des boissons alcoolisées, des vernis
    pour céramique et de la verrerie de cristal. Dans les pays oů l'on
    utilise encore de l'essence au plomb, les principales émissions dans
    l'atmosphčre proviennent de sources de combustion de produits
    pétroliers, sources qui peuvent ętre mobiles ou fixes (en
    agglomération). Les émissions dans l'atmosphčre sont également trčs
    importantes ŕ proximité des mines et des fonderies de plomb.

         Le plomb en suspension dans l'air peut se déposer sur le sol et
    dans l'eau et il parvient ainsi jusqu'ŕ l'homme par l'intermédiaire de
    la chaîne alimentaire et de l'eau de boisson. Le plomb atmosphérique
    contribue également pour une part importante au plomb présent dans la
    poussičre domestique.

    3.  Transport, distribution et transformation dans l'environnement

         Le transport et la distribution du plomb ŕ partir de sources
    fixes, mobiles ou naturelles, s'effectue principalement par
    l'intermédiaire de l'air. La majeure partie du plomb émis dans l'air
    se redépose ŕ proximité de la source d'émission, męme si quelques
    particules (de diamčtre < 2 µm) dont transportées sur de longues
    distances et provoquent la contamination de sites aussi reculés que
    les glaciers arctiques. Le plomb aéroporté peut contribuer ŕ
    l'exposition humaine par contamination des aliments, d l'eau et de la
    poussičre, ainsi que par inhalation directe. L'élimination du plomb
    aéroporté dépend des conditions atmosphériques et de la taille des
    particules en suspension. De grandes quantités de plomb peuvent ętre
    déversées dans le sol et l'eau. Cependant, ces produits ont tendance ŕ
    rester sur place du fait de la médiocre solubilité dans l'eau des
    dérivés du plomb.

         Le plomb qui s'est déposé dans l'eau, que ce soit ŕ partir de
    l'air ou par lessivage des sols, se répartit rapidement entre les
    sédiments et la phase aqueuse, selon la valeur du pH, la teneur en
    sels et la présence éventuelle d'agents chélateurs organiques. Lorsque
    le pH dépasse 5,4, l'eau dure peu contenir environ 30 µg de
    plomb/litre et l'eau douce environ 500 µg/litre. Une trčs faible

    quantité du plomb déposé sur le sol passe dans les eaux de surface ou
    les eaux souterraines sauf en cas d'érosion ou d'altération des
    roches; il est normalement fortement lié par chélation aux matičres
    organiques.

         Le plomb en suspension dans l'air peut passer dans les ętres
    vivants soit directement, soit par fixation ŕ partir du sol. Les
    animaux peuvent ętre exposés au plomb de maničre directe par broutage
    de l'herbe, ingestion de terre ou par inhalation. Le plomb minéral ne
    subit qu'une faible bioamplification le long de la chaîne alimentaire.

    4.  Concentrations dans l'environnement et exposition humaine

         Dans la population générale adulte des non-fumeurs, la principale
    voie d'exposition est la consommation de nourriture et d'eau. Le plomb
    est suspension dans l'air peut jouer un rôle important dans
    l'exposition, en fonction de facteurs tels que le tabagisme, la
    profession, la proximité d'une autoroute, d'une fonderie, etc. ou de
    certains lieux de loisirs (par exemple, ateliers d'artisanat, stands
    de tir, etc.). Pour les nourrissons et les enfants en bas âge, la
    nourriture, l'air, l'eau et la poussičre ou la terre sont les
    principales voies d'exposition potentielles. Pour les nourrissons
    jusqu'ŕ quatre ou cinq mois, les principales sources d'exposition au
    plomb sont l'air, le lait et les laits maternisés ainsi que l'eau.

         Les concentrations de plomb observées dans l'air, les aliments,
    l'eau et le sol ou la poussičre varient largement d'une région ŕ
    l'autre du monde et dépendent du degré d'industrialisation,
    d'urbanisation ainsi que du mode de vie. Dans l'air ambiant, des
    concentrations supérieures ŕ 10 µg/m3 ont été signalées en milieu
    urbain ŕ proximité d'une fonderie, alors que dans les villes oů l'on
    n'utilise plus d'essence au plomb, ces concentrations peuvent tomber
    en-dessous de 0,2 µg/m3. Dans ces conditions, la dose de plomb
    absorbée ŕ partir de l'air peut varier de moins de 4 µg/jour ŕ plus de
    200 µg/jour.

         Dans des échantillons d'eau potable prélevés ŕ la source, on
    trouve généralement des concentrations de plomb inférieures ŕ
    5 µg/litre. Cependant, lorsque l'eau est prélevée au robinet dans les
    maisons dont les canalisations sont en plomb, les teneurs peuvent
    dépasser 100 µg/litre, en particulier lorsque l'eau a séjourné dans la
    tuyauterie pendant plusieurs heures.

         L'exposition au plomb par la voie alimentaire dépend de nombreux
    facteurs tenant au mode de vie et notamment ŕ la nature des aliments
    consommés, au mode de préparation, au fait que l'on utilise ou non de
    la brasure au plomb, ŕ la teneur de l'eau en plomb et ŕ l'utilisation
    de vaisselle recouverte d'un vernis au plomb.

         Les nourrissons et les enfants peuvent souvent ętre trčs exposés
    au plomb présent dans la poussičre et dans la terre. La teneur de la
    poussičre en plomb dépend de facteurs tels que la vétusté et l'état de
    la maison, l'utilisation de peintures ŕ base de plomb, la présence de
    plomb dans l'essence et la densité urbaine. La dose de plomb absorbée
    dépendra également de l'âge et du comportement de l'enfant ainsi que
    de la biodisponibilité du plomb dans le produit oů il est présent.

         Chez les ouvriers des industries oů l'on produit, raffine,
    utilise ou rejette du plomb ou des dérivés du plomb, la principale
    voie d'exposition est l'inhalation. Au cours d'un poste de travail de
    8 heures, les ouvriers peuvent en absorber des quantités atteignant
    400 µg, qui s'ajoutent aux 20 ou 30 µg absorbés quotidiennement ŕ
    partir de la nourriture, de l'eau et de l'air ambiant; l'ingestion de
    grosses particules peut également contribuer de façon importante ŕ cet
    apport.

    5.  Cinétique et métabolisme chez les animaux de laboratoire et
        l'homme

         L'homme et les animaux résorbent le plomb qu'ils ingčrent ou
    inhalent; la résorption percutanée est minime chez l'homme. Selon sa
    granulométrie, sa forme chimique et sa solubilité dans les liquides
    biologiques, un dérivé du plomb peut ętre résorbé jusqu'ŕ hauteur de
    50% aprčs avoir été inhalé. Certaines grosses particules (plus de
    7 µm) sont avalées aprčs avoir été rejetées des voies respiratoires
    par l'action de l'ascenseur muco-ciliaire. Chez l'homme et les animaux
    d'expérience, l'absorption du plomb dans les voies digestives dépend
    de la nature physico-chimique du produit ingéré, de l'état
    nutritionnel du sujet ainsi que du type d'aliments consommés. Chez
    l'homme adulte, le plomb contenu dans les aliments est absorbé ŕ peu
    prčs ŕ hauteur de 10%; la proportion est plus élevée lorsque le sujet
    est ŕ la dičte. Toutefois, chez les nourrissons et les enfants en bas
    âge, le plomb d'origine alimentaire peut ętre absorbé ŕ hauteur de
    50%, encore que le taux d'absorption du plomb provenant des poussičres
    ou de la terre ainsi que des écailles de peinture puisse ętre plus
    faible, selon sa biodisponibilité. Les régimes alimentaires pauvres en
    calcium, phosphates, sélénium ou zinc peuvent accroître l'absorption
    du plomb. Le fer et la vitamine D influent également sur l'absorption
    du plomb.

         Le taux de plomb dans le sang ou plombémie est utilisé pour
    évaluer la charge en plomb de l'organisme ainsi que les doses de plomb
    (internes) absorbées. La relation entre la plombémie et la
    concentration du plomb dans les diverses sources d'exposition n'est
    pas linéaire (corrélation curviligne).

         Une fois absorbé, le plomb ne se répartit pas de maničre homogčne
    dans l'organisme. Aprčs ętre rapidement passé dans le sang et les
    tissus mous, il se redistribue lentement des les os. La plomb
    s'accumule dans les os pendant une longue période de la vie humaine et

    peut servir de source endogčne de plomb. La demi-vie du plomb dans le
    sang et les tissus mous est d'environ 28 ŕ 36 jours mais elle peut
    ętre beaucoup plus longue dans les diverses parties de l'os. Le taux
    de rétention du plomb dans l'organisme est plus élevé chez l'enfant
    que chez l'adulte. Pendant toute la durée de la gestation, le plomb
    passe facilement de la mčre au foetus.

         La plombémie est la mesure la plus couramment utilisée pour
    évaluer l'exposition au plomb. Toutefois, on dispose aujourd'hui de
    techniques permettant de doser le plomb dans les dents et les os,
    encore que la cinétique du phénomčne ne soit pas parfaitement
    élucidée.

    6.  Effets sur les animaux de laboratoire et les systčmes d'épreuves
        in vitro

         Chez toutes les espčces d'animaux de laboratoire étudiées, y
    compris des primates non-humains, on a constaté que le plomb
    produisait des effets indésirables au niveau de plusieurs organes et
    systčmes, notamment le systčme haematopoďétique, le systčme nerveux,
    les reins, le systčme cardio-vasculaire, l'appareil reproducteur et le
    systčme immunitaire. Le plomb est également nocif pour les os et on a
    montré qu'il avait des effets cancérogčnes sur le rat et la souris.

         Malgré les différences d'ordre cinétique qui existent entre les
    animaux d'expérience et l'homme, ces études apportent des arguments
    biologiques de poids ŕ la plausibilité de tels effets chez l'homme.
    Chez le rat, on a observé une diminution de la capacité
    d'apprentissage et de mémorisation lorsque la plombémie était de
    l'ordre de 0,72 ŕ 0,96 µmol/litre (soit 15 ŕ 20 µg/dl), les męmes
    effets étant observés chez des primates non-humains pour une plombémie
    ne dépassant pas 0,72 µmol/litre (15 µg/dl). En outre,
    l'expérimentation animale a également permis d'observer des troubles
    de la vision et de l'audition.

         Chez le rat, une plombémie supérieure ŕ 2,88 µmol/litre
    (60 µg/dl), soit une valeur analogue ŕ celle qui, selon les
    observations, constitue le seuil d'apparition des effets chez l'homme,
    est ŕ męme de provoquer l'apparition d'effets néphrotoxiques. Des
    effets cardio-vasculaires ont été également observés chez des rats
    aprčs exposition chronique ŕ de faibles doses de plomb entraînant une
    plombémie de l'ordre de 0,24 ŕ 1,92 µmol/litre (5 ŕ 40 µg/dl). A des
    doses inférieures ŕ la dose maximale tolérée qui est de 200 mg de
    plomb (sous forme d'acétate) par litre d'eau potable, on a observé
    l'apparition de tumeurs. Cette dose constitue la dose maximale qui
    n'entraîne pas d'autres effets morphologiques ou fonctionnels.

    7.  Effets sur l'homme

         Chez l'homme, le plomb peut produire des effets biologiques trčs
    divers selon l'intensité et la durée de l'exposition. On a ainsi
    observé des effets au niveau infracellulaire ainsi que des effets
    s'exerçant sur les fonctions générales de l'organisme, effets qui vont
    de l'inhibition de certaines enzymes jusqu'ŕ l'apparition
    d'altérations morphologiques marquées et ŕ la mort. Ces altérations se
    produisent dans de larges limites de doses, l'organisme humain en
    développement étant généralement plus sensible que l'organisme adulte.

         On a montré que le plomb avait des effets sur nombre de processus
    biochimiques; en particulier on a largement étudié les effets qu'il
    produit sur la synthčse hémique, tant chez l'adulte que chez l'enfant.
    Lorsque la plombémie est élevée, on observe une augmentation du taux
    de protoporphyrine érythrocytaire sérique ainsi qu'une augmentation de
    l'excrétion urinaire de coproporphyrine et d'acide delta-aminolévulinique.
    A des valeurs plus faibles de la plombémie, on observe l'inhibition
    d'enzymes comme la delta-aminolévulinique acide-déshydratase et la
    dihydrobioptérine-réductase.

         Les effets du plomb sur le systčme haématopoďétique se traduisent
    par une diminution de la synthčse de l'hémoglobine et l'on a observé
    une anémie chez des enfants lorsque la plombémie dépassait
    1,92 µmol/litre (40 µg/dl).

         Pour des raisons d'ordre neurologique, métabolique et
    compartementale, les enfants sont plus sensibles aux effets du plomb
    que les adultes. Des études épidémiologiques tant prospectives que
    transversales ont été effectuées pour déterminer dans quelle mesure
    une exposition au plomb présent dans l'environnement affecte les
    fonctions psychologiques dépendant du systčme nerveux central. On a
    ainsi montré qu'il existait une association entre l'exposition au
    plomb et les troubles des fonctions neurocomportementales chez
    l'enfant.

         On a également constaté des anomalies des fonctions
    psychologiques et neurocomportementales chez des ouvriers longtemps
    exposés au plomb. Il a été montré que les paramčtres
    électrophysiologiques étaient d'utiles indicateurs des effets
    infracliniques du plomb sur le systčme nerveux central.

         On sait depuis longtemps que la neuropathie périphérique est due
    ŕ une exposition prolongée ŕ de fortes concentrations de plomb sur le
    lieu de travail. A plus faibles concentrations, on a observé une
    diminution de la vitesse de conduction nerveuse. Ces effets se
    révčlent souvent réversibles, en fonction de l'âge et de la durée de
    l'exposition, aprčs cessation de l'exposition.

         Les effets que le plomb exerce sur le coeur sont indirects et
    s'opčrent par l'intermédiaire du systčme neuro-végétatif; il n'y a pas
    d'effets directs sur le myocarde. L'ensemble des faits tirés des
    études sur des populations d'adultes indiquent qu'il existe une trčs
    faible association entre la plombémie et la tension artérielle
    systolique ou diastolique. Etant donné la difficulté qu'il y a ŕ tenir
    compte des facteurs de confusion, il n'a pas été possible d'établir, ŕ
    partir des résultats de ces études, une relation de cause ŕ effet.
    Rien n'indique non plus que l'association qui pourrait exister entre
    la plombémie et la tension artérielle constitue un problčme médical
    majeur.

         On sait que le plomb peut entraîner des lésions tubulaires
    proximales qui se caractérisent par une aminoacidurie généralisée, une
    hypophosphatémie avec d'une hyperphosphaturie relative et une
    glycosurie accompagnée d'inclusions nucléiques, d'altérations des
    mitochondries et d'une hypertrophie des cellules de l'épithélium
    tubulaire proximal. Ces effets sur les tubules s'observent aprčs une
    exposition relativement brčve et sont généralement réversibles. En
    revanche, les altérations scléreuses et les fibroses interstitielles
    qu'on observe lors d'une exposition chronique ŕ de fortes
    concentrations de plomb, entraînent des problčmes fonctionnels qui
    peuvent déboucher sur une insuffisance rénale. On a observé une
    augmentation du risque de néphropathie chez les travailleurs dont la
    plombémie dépassait 3,0 µmol/litre (soit environ 60 µg/dl). En
    utilisant des indicateurs fonctionnels plus sensibles, on a récemment
    découvert l'existence d'effets rénaux dans la population générale.

         Les effets du plomb sur la fonction de reproduction, ne
    concernent, chez l'homme, que la morphologie et le nombre des
    spermatozoďdes. Chez la femme, on a attribué au plomb un certain
    nombre de grossesses ŕ issue défavorable.

         Il ne semble pas que le plomb ait des effets nocifs sur la peau,
    les muscles ou le systčme immunitaire. Si l'on excepte le cas du rat,
    le plomb ne paraît pas non plus ętre ŕ l'origine de l'apparition de
    tumeurs.

    8.  Evaluation des risques pour la santé humaine

         Le plomb a des effets nocifs sur plusieurs organes et systčmes,
    les effets les plus sensibles étant relevés au niveau infracellulaire
    ainsi que sur le développement du systčme nerveux. On a fait état
    d'une association entre la plombémie et l'hypertension. En outre, le
    plomb entraîner une cascade d'effets sur les réserves hémiques de
    l'organisme et il perturbe également la synthčse hémique. Cependant,
    certains de ces effets ne sont pas véritablement considérés comme
    délétčres. Il y a également perturbation de l'homéostase du calcium
    avec des contre-coups sur d'autres processus cellulaires.

    a)   Les preuves les plus substantielles sont on dispose ŕ propos de
         l'action nocive du plomb proviennent d'études transversales et
         prospectives sur des populations dont la plombémie est
         généralement inférieure ŕ 1,2 µmol/litre (25 µg/dl) et concernent
         une diminution du quotient d'intelligence (QI). Il importe
         cependant de noter que ces observations ne constituent pas une
         preuve définitive d'une relation de cause ŕ effet entre ce
         phénomčne et l'exposition au plomb. Cependant la mesure de ce QI
         ŕ partir de l'âge de 4 ans fait ressortir un déficit qui se situe
         entre 0 et 5 points (sur une échelle avec un écart-type de 15)
         pour une augmentation de la plombémie de 0,48 µmol/litre
         (10 µg/dl), l'ampleur de l'effet étant vraisemblablement de 1 ŕ 3
         points. Lorsque la plombémie dépasse 1,2 µmol/litre (25 µg/dl),
         on peut avoir une relation différente entre cette plombémie et le
         QI. Les estimations de l'ampleur de l'effet observé sont des
         moyennes calculées sur des groupes et ne représentent qu'une
         probabilité pour un individu en particulier.

              Les études épidémiologiques dont on dispose ne donnent pas
         de preuves définitives de l'existence d'un seuil. Lorsque la
         plombémie se situe en-dessous de l'intervalle
         0,48-0,72 µmol/litre (10-15 µg/dl), l'incertitude qui entache
         toute estimation de l'effet s'accroît, du fait des facteurs de
         confusion et des limites dans la précision des dosages et des
         tests psychométriques. Il n'en reste pas moins qu'en-dessous de
         cet intervalle de valeurs, certains éléments incitent ŕ penser ŕ
         l'existence d'une association.

    b)   L'expérimentation animale milite en faveur de l'existence d'une
         relation de cause ŕ effet entre l'exposition au plomb et certains
         effets neurologiques, puisqu'elle fait ressortir des déficits
         dans les fonctions cognitives pour une plombémie ne dépassant pas
         0,53 ŕ 0,72 µmol/litre (11-15 µg/dl), déficits qui peuvent
         persister bien aprčs la cessation de l'exposition au plomb.

    c)   Pour une plombémie ne dépassant pas 1,44 µmol/litre (30 µg/dl),
         il peut y avoir réduction de la vitesse de conduction nerveuse
         périphérique chez l'homme. En outre, pour des  valeurs de la
         plombémie de dépassant pas 1,92 µmol/litre (40 µg/dl), il peut
         également y avoir perturbation des fonctions moto-sensorielles et
         la fonction du systčme nerveux neurovégétatif (variabilité de
         l'intervalle R-R sur l'électrocardiogramme) peut ętre affectée
         pour une valeur moyenne de la plombémie d'environ 1,68 µmol/litre
         (35 µg/dl). Chez les travailleurs dont la plombémie dépasse
         2,88 µmol/litre (60 µg/dl), il y a accroissement du risque de
         néphropathie saturnienne. Toutefois, des études récentes basées
         sur des indicateurs plus sensibles de la fonction rénale incitent
         ŕ penser que des effets peuvent se produire ŕ des valeurs plus
         faibles de l'exposition au plomb.

    d)   Il semblerait que l'exposition au plomb soit associée ŕ une
         légčre augmentation de la tension artérielle. L'ordre de grandeur
         probable de cette augmentation est le suivant: pour un doublement
         de la plombémie (par exemple lorsqu'elle passe de 0,8 ŕ
         1,6 µmol/litre, soit de 16,6 ŕ 33,3 µg/dl), il y a une
         augmentation moyenne de 1 mmHg de la systolique. L'association
         avec la diastolique est analogue mais d'ampleur plus faible.
         Toutefois, on se demande si ces associations statistiques
         résultent réellement de l'exposition au plomb ou s'il s'agit d'un
         artefact imputable ŕ des facteurs de confusion.

    e)   Certaines études épidémiologiques - mais pas toutes - font état
         d'une association liée ŕ la dose entre les accouchements avant
         terme et certains indices de la croissance et de la maturation
         foetales ŕ des valeurs de la plombémie supérieures ou égales ŕ
         0.72 µmol/litre (15 µg/dl).

    f)   Les données relatives ŕ la cancérogénicité pour l'homme du plomb
         et de plusieurs de ses dérivés minéraux, sont insuffisantes.

    g)   On a mis en évidence des effets que le plomb exerce sur un
         certain nombre de systčmes enzymatiques et de paramčtres
         biochimiques. Les valeurs de la plombémie au-dessus desquelles on
         peut mettre en évidence des effets avec les techniques actuelles,
         pour ce qui est des paramčtres susceptibles d'avoir une
         importance clinique, sont toutes supérieures ŕ 0,96 µmol/litre
         (20 µg/dl). Certains effets sur les enzymes peuvent ętre mis en
         évidence ŕ des valeurs plus faibles de la plombémie, mais leur
         signification clinique demeure incertaine.

    RESUMEN

         La presente monografía se centra en los riesgos para la salud
    humana asociados a la exposición al plomo y a los compuestos
    inorgánicos de plomo. Se han destacado los datos disponibles desde la
    publicación de Criterios de Salud Ambiental No. 3: Plomo (OMS, 1977).
    Los efectos ambientales del plomo se examinan en Environmental Health
    Criteria 85: Lead - Environmental Aspects (OMS, 1989).

    1.  Identidad, propiedades físicas y químicas y métodos analíticos

         El plomo es un metal blando, gris plateado, que se funde a
    327,5°C. Es muy resistente a la corrosión, pero es soluble en ácido
    nítrico y en ácido sulfúrico caliente. Su valencia corriente en los
    compuestos inorgánicos es +2. Su solubilidad en agua varía; el sulfito
    de plomo y los óxidos de plomo son poco solubles, mientras que las
    sales de nitrato, clorato y cloruro son razonablemente solubles en
    agua fría. El plomo también forma sales con ácidos orgánicos tales
    como el láctico y el acético, y compuestos orgánicos estables tales
    como el tetraetilo de plomo y el tetrametilo de plomo.

         Los métodos utilizados más corrientemente para el análisis de
    bajas concentraciones de plomo en materias biológicas y ambientales
    son la llama, el horno de grafito y la espectroscopia de absorción
    atómica de plasma acoplado inductivamente y la voltimetría de
    separación anódica. Según sean el tratamiento previo de la muestra,
    las técnicas de extracción y la instrumentación analítica, pueden
    alcanzarse niveles de detección de 0,12 µmoles de plomo por litro de
    sangre (2,49 µg/dl). Sin embargo, se obtienen resultados fiables sólo
    cuando se siguen procedimientos específicos para reducir al mínimo el
    riesgo de contaminación durante la recogida, el almacenamiento,
    procesamiento y análisis de la muestra.

    2.  Fuentes de exposición humana

         El nivel de plomo en la corteza terrestre es de aproximadamente
    20 mg/kg. El plomo del medio ambiente puede provenir de fuentes
    naturales o antropogénicas. Las fuentes naturales de plomo atmosférico
    comprenden el desgaste geológico y las emisiones volcánicas y se han
    estimado en 19 000 toneladas por ańo, frente a unas 126 000 toneladas
    por ańo emitidas en el aire como resultado de la minería, la fundición
    y el consumo de más de 3 millones de toneladas de plomo por ańo.

         Se han encontrado concentraciones atmosféricas de plomo de
    50 pg/m3 en zonas remotas. Los niveles básicos de plomo en el suelo
    oscilan entre 10 y 70 mg/kg y se ha comunicado un nivel medio de
    138 mg/kg en las proximidades de las carreteras. Los niveles de plomo
    presentes en las aguas rara vez exceden de unos pocos microgramos por
    litro; la concentración natural de plomo en las aguas superficiales se
    ha estimado en 0,02 µg/litro.

         El plomo y sus compuestos pueden entrar en el medio ambiente en
    cualquier punto durante las actividades de minería, fundición,
    elaboración, utilización, reciclado o eliminación. Se utiliza
    principalmente en la fabricación de pilas, cables, pigmentos, aditivos
    de la gasolina, productos para soldar y de acero. El plomo y los
    compuestos de plomo también se utilizan para soldar las tuberías de
    distribución de agua y las latas de conserva, en algunos remedios
    tradicionales, en las tapas de las botellas de bebidas alcohólicas y
    en los esmaltes cerámicos y la cristalería de mesa. En los países
    donde todavía se utiliza gasolina con plomo, la principal emisión en
    el aire proviene de fuentes móviles y estacionarias de combustión de
    gasolina (centros urbanos). Las zonas próximas a las minas y funderías
    de plomo están expuestas a la emisión de niveles elevados en el aire.

         El plomo del aire puede depositarse en el suelo y el agua, desde
    donde llega al ser humano por conducto de la cadena alimentaria y del
    agua de bebida. El plomo atmosférico también es una fuente importante
    del plomo presente en el polvo de las viviendas.

    3.  Transporte, distribución y transformación en el medio ambiente

         El transporte y la distribución del plomo procedente de fuentes
    fijas, móviles y naturales tienen lugar principalmente a través del
    aire. La mayor parte de las emisiones de plomo se depositan cerca de
    la fuente, aunque algunas partículas de materia (< 2 µm de diámetro)
    recorren largas distancias y contaminan lugares remotos tales como los
    glaciares árticos. El plomo del aire puede contribuir a la exposición
    humana mediante la contaminación de los alimentos, del agua y del
    polvo, así como por inhalación directa. La eliminación del plomo del
    aire depende de las condiciones atmosféricas y del tamańo de las
    partículas. Pueden descargarse grandes cantidades de plomo en el suelo
    y en el agua. Sin embargo, ese material tiende a permanecer localizado
    debido a la escasa solubilidad de los compuestos de plomo en el agua.

         El plomo depositado en el agua, ya provenga del aire o de la
    escorrentía del suelo, se distribuye rápidamente entre el sedimento y
    la fase acuosa, según el pH, el contenido de sales y la presencia de
    agentes quelantes orgánicos. Con un pH superior a 5,4, las aguas duras
    pueden contener aproximadamente 30 µg de plomo por litro y las aguas
    blandas aproximadamente 500 µg de plomo por litro. Muy poco plomo
    depositado en el suelo se transporta a las aguas superficiales o a las
    subterráneas, salvo mediante la erosión o el desgaste geoquímico;
    normalmente está ligado a la materia orgánica de forma bastante
    estrecha (por quelación).

         El plomo del aire puede transferirse a la biota directamente o
    por absorción del suelo. Los animales pueden encontrarse expuestos al
    plomo directamente mediante la ingestión de hierba y de tierra o por
    inhalación. Hay poca biomagnificación del plomo inorgánico a través de
    la cadena alimentaria.

    4.  Niveles ambientales y exposición humana

         En la población general que no fuma, la principal vía de
    exposición son los alimentos y el agua. El plomo del aire puede
    contribuir apreciablemente a la exposición, lo que depende de factores
    tales como el consumo de tabaco, la ocupación, la proximidad de
    caminos transitados por vehículos automotores, de funderías de plomo,
    etc., y ciertas actividades de esparcimiento (por ejemplo, artesanía,
    tiro con armas de fuego). Los alimentos, el aire, el agua y el
    polvo/suelo son las principales vías potenciales de exposición de los
    nińos pequeńos. Para los nińos de hasta 4 ó 5 meses de edad, el aire,
    la leche, las preparaciones para lactantes y el agua son fuentes
    notables de exposición al plomo.

         Los niveles de plomo presentes en el aire, los alimentos, el
    agua, y el suelo/polvo varían ampliamente en el mundo y dependen del
    grado de desarrollo industrial y de urbanización y de factores
    relacionados con el modo de vida. Se han comunicado niveles superiores
    a 10 µg/m3 presentes en el aire ambiental en zonas urbanas próximas
    a funderías, mientras que en ciudades donde ha dejado de usarse la
    gasolina con plomo se han detectado niveles inferiores a 0,2 µg/m3.
    La absorción de plomo del aire puede, pues, variar de menos de
    4 µg/día a más de 200 µg/día.

         Los niveles de plomo en muestras de agua de bebida extraídas de
    los manantiales suelen ser inferiores a 5 µg/litro. Sin embargo, el
    agua del grifo de viviendas cuyas tuberías tienen plomo contiene
    niveles que exceden de 100 µg/litro, en particular después de haber
    reposado el agua en las tuberías durante algunas horas.

         El nivel de exposición al plomo a través de la dieta depende de
    muchos factores relacionados con el modo de vida, entre ellos los
    alimentos que se consumen, la tecnología de elaboración, el empleo de
    soldadura de plomo, los niveles de plomo en el agua y la utilización
    de cerámica con barniz de plomo.

         Para los nińos, el plomo presente en el polvo y en el suelo suele
    ser la principal vía de exposición. Los niveles de plomo en el polvo
    dependen de factores tales como la antigüedad y el estado de la
    vivienda, la utilización de pinturas a base de plomo, el plomo de la
    gasolina y la densidad urbana. La absorción de plomo dependerá de la
    edad y de las características comportamentales del nińo así como de la
    biodisponibilidad de plomo en la fuente material.

         La inhalación es la vía principal de exposición al plomo para los
    trabajadores de industrias que producen, refinan, utilizan o desechan
    plomo y compuestos de plomo. Durante un turno de ocho horas, los
    trabajadores pueden absorber nada menos que 400 µg de plomo, además de
    los 20-30 µg/día que absorben de los alimentos, del agua y del aire
    ambiental; puede haber una absorción notable como resultado de la
    inhalación de partículas grandes.

    5.  Cinética y metabolismo en animales de laboratorio y en el ser
        humano

         Los seres humanos y los animales absorben plomo por inhalación o
    por ingestión; la absorción percutánea es mínima en el ser humano.
    Según la especiación química, el tamańo de las partículas y la
    solubilidad de los líquidos corporales, puede absorberse hasta un 50%
    de los compuestos de plomo inhalados. Algunas partículas de materia
    inhaladas (de más de 7 µm) se degluten después de la eliminación
    mucociliar del aparato respiratorio. En los animales experimentales y
    en el ser humano, la absorción de plomo del aparato gastrointestinal
    está influenciada por la naturaleza fisicoquímica del material
    ingerido, el estado nutricional y el tipo de alimentación. En los
    seres humanos adultos, se absorbe aproximadamente el 10% del plomo
    contenido en la alimentación; la proporción es más elevada en
    condiciones de ayuno. Sin embargo, los lactantes y los nińos pequeńos
    absorben nada menos que el 50% del plomo presente en la alimentación,
    pero la absorción de plomo del polvo/suelo y de desconchones de
    pintura puede ser menor y depende de su biodisponibilidad. Las dietas
    pobres en calcio, fosfato, selenio o zinc pueden dar lugar a una mayor
    absorción de plomo. El hierro y la vitamina D también influyen en la
    absorción de plomo.

         Los niveles de plomo en la sangre (Pb-H) se utilizan para medir
    la carga corporal y las dosis absorbidas (internas) de plomo. La
    relación entre el plomo presente en la sangre y la concentración de
    plomo en las fuentes de exposición es curvilínea.

         Una vez absorbido, el plomo no se distribuye de manera homogénea
    en todo el cuerpo. Hay una absorción rápida en la sangre y en los
    tejidos blandos, seguida de una redistribución más lenta a los huesos.
    Los huesos acumulan plomo durante gran parte de la vida humana y
    pueden actuar como fuente endógena de plomo. La semivida del plomo en
    la sangre y en otros tejidos blandos es de aproximadamente 28-36 días,
    pero es mucho más larga en los diversos compartimentos óseos. La
    retención porcentual de plomo en los depósitos corporales es más
    elevada en los nińos que en los adultos. La transferencia de plomo al
    feto humano se efectúa fácilmente durante la gestación.

         El nivel de plomo en la sangre es la medida más utilizada para
    determinar la exposición al plomo. Sin embargo, ya se dispone de
    técnicas para determinar la cantidad de plomo presente en los dientes
    y en los huesos, aunque aún no se conoce del todo su cinética.

    6.  Efectos en los animales de laboratorio y en los sistemas in vitro

         En todas las especies de animales de experimentación estudiadas,
    inclusive en primates no humanos, se ha observado que el plomo tiene
    efectos adversos en varios órganos y sistemas de órganos, inclusive

    los sistemas hematopoyético, nervioso, renal, cardiovascular,
    reproductivo e inmunitario. El plomo también afecta a los huesos y se
    ha demostrado que es carcinógeno en ratas y ratones.

         Pese a diferencias cinéticas entre las especies de animales
    experimentales y la humana, dichos estudios apoyan firmemente y
    justifican desde un punto de vista biológico los hallazgos realizados
    en seres humanos. Se han comunicado deficiencias del aprendizaje y de
    la memoria en ratas con niveles de Pb-H de 0,72-0,96 µmoles/litro
    (15-20 µg/dl) y en primates no humanos con niveles de Pb-H de no más
    de 0,72 µmoles/litro (15 µg/dl). Además, se han comunicado
    deficiencias visuales y auditivas en estudios realizados en animales
    de experimentación.

         La toxicidad renal en las ratas parece presentarse a partir de un
    nivel de Pb-H de 2,88 µmoles/litro (60 µg/dl); este valor es semejante
    al que, según se ha comunicado, comienza a tener efectos renales en el
    ser humano. Se han observado efectos cardio-vasculares en ratas
    después de exposiciones crónicas a niveles bajos que dan lugar a
    niveles de Pb-H de 0,24-1,92 µmoles/litro (5-40 µg/dl). Se ha
    demostrado que aparecen tumores con dosis inferiores a la dosis máxima
    tolerada, de 200 mg de plomo (como acetato de plomo) por litro de agua
    de bebida. Esta es la dosis máxima no asociada a otros cambios
    morfológicos o funcionales.

    7.  Efectos en el ser humano

         En el ser humano, el plomo puede tener una amplia variedad de
    efectos biológicos según el nivel y la duración de la exposición. Se
    han observado efectos en el plano subcelular y efectos en el
    funcionamiento general del organismo que van desde la inhibición de
    las enzimas hasta la producción de acusados cambios morfológicos y la
    muerte. Dichos cambios se producen a dosis muy diferentes; en general,
    el ser humano que se está desarrollando es más sensible que el adulto.

         Se ha mostrado que el plomo tiene efectos en muchos procesos
    bioquímicos; en particular, se han estudiado mucho los efectos en la
    síntesis del hemo en adultos y nińos. Se observan niveles más altos de
    porfirina eritrocitaria sérica y mayor excreción urinaria de
    coproporfirina y de ácido delta-aminolevulínico cuando las
    concentraciones de Pb-H son elevadas. Con niveles más bajos se observa
    inhibición de las enzimas dehidratasa del ácido delta-aminolevulínico
    y reductasa de la dihidrobiopterina.

         Como resultado de los efectos del plomo en el sistema
    hematopoyético disminuye la síntesis de hemoglobina y se ha observado
    anemia en nińos a concentraciones de Pb-H superiores a
    1,92 µmoles/litro (40 µg/dl).

         Por razones neurológicas, metabólicas y comportamentales, los
    nińos son más vulnerables a los efectos del plomo que los adultos. Se
    han efectuado estudios epidemiológicos prospectivos y transversales
    para evaluar la medida en que la exposición al plomo ambiental afecta
    a las funciones psicológicas regidas por el sistema nervioso central
    (SNC). Se ha mostrado que el plomo está asociado a deficiencias
    neurocomportamentales en los nińos.

         Se han observado deficiencias psicológicas y
    neurocomportamentales en trabajadores que habían estado expuestos al
    plomo durante un tiempo prolongado. Los parámetros electrofisiológicos
    han demostrado ser indicadores útiles de los efectos subclínicos del
    plomo en el SNC.

         Desde hace tiempo se sabe que la exposición prolongada a niveles
    elevados de plomo en el medio laboral provoca neuropatías periféricas.
    Con niveles más bajos se ha observado una reducción de la velocidad de
    conducción nerviosa. Se ha observado a menudo que dichos efectos son
    reversibles después de cesar la exposición, según la edad del sujeto y
    la duración de la exposición.

         Los efectos del plomo en el corazón son indirectos y se producen
    por conducto del sistema nervioso autónomo; el plomo no tiene efectos
    directos en el miocardio. Datos colectivos procedentes de estudios de
    poblaciones adultas indican asociaciones muy débiles entre la
    concentración de Pb-H y la presión arterial sistólica o diastólica.
    Dada la dificultad de evaluar el influjo de los factores de confusión
    pertinentes, no puede establecerse una relación causal sobre la base
    de esos estudios. No hay indicios de que la relación entre la
    concentración de Pb-H y la presión arterial tenga mucha importancia
    para la salud.

         Se sabe que el plomo provoca en los tubos proximales del rińón
    lesiones que se caracterizan por aminoaciduria generalizada,
    hipofosfatemia con hiperfosfaturia relativa y glucosuria acompańada de
    cuerpos de inclusión nuclear, modificaciones mitocondriales y
    citomegalia de las células epiteliales de los tubos proximales. Los
    efectos tubulares se manifiestan después de una exposición
    relativamente breve y suelen ser reversibles, mientras que los cambios
    escleróticos y la fibrosis intersticial, que dan lugar a una
    disminución de la función renal y a una posible insuficiencia renal,
    requieren una exposición crónica a niveles elevados de plomo. Se ha
    advertido un mayor riesgo de nefropatía en los trabajadores que tienen
    niveles de Pb-H superiores a 3,0 µmoles/litro (aproximadamente
    60 µg/dl). Recientemente se han observado efectos renales en la
    población general tras haberse utilizado indicadores de función más
    sensibles.

         Los efectos del plomo en la función reproductora masculina se
    limitan a la morfología y el número de los espermatozoides. En cuanto
    a la femenina, se han atribuido al plomo algunos efectos adversos en
    el embarazo.

         El plomo no parece tener efectos nocivos en la piel, en los
    músculos ni en el sistema inmunitario. Salvo en la rata, el plomo no
    parece estar relacionado con el desarrollo de tumores.

    8.  Evaluación de los riesgos para la salud humana

         El plomo tiene efectos adversos en varios órganos y sistemas de
    órganos; los más delicados parecen ser los cambios subcelulares y los
    efectos en el desarrollo del sistema nervioso. Se ha observado una
    asociación entre el nivel de Pb-H y la hipertensión (presión
    arterial). El plomo produce una serie de efectos en la reserva
    corporal de hemo y afecta a la síntesis de éste. Sin embargo, algunos
    de estos efectos no se consideran adversos. Está afectada la
    homeostasia del calcio, lo que interfiere en otros procesos celulares.

    a)   Los datos más importantes de los estudios transversales y
         prospectivos de poblaciones con niveles de Pb-H generalmente
         inferiores a 1,2 µmoles/litro (25 µg/dl) se relacionan con una
         disminución del coeficiente de inteligencia (CI). Es importante
         seńalar que esas observaciones no pueden constituir una prueba
         concluyente de una relación causal con la exposición al plomo.
         Sin embargo, la magnitud del efecto aparente en el CI,
         determinado desde los 4 ańos en adelante, es un déficit de 0 a
         5 puntos (en una escala con una desviación estándar de 15) por
         cada 0,48 µmoles/litro (10 µg/dl) de aumento del nivel de Pb-H,
         con una magnitud probable del efecto aparente de 1 a 3 puntos.
         Con niveles de Pb-H superiores a 1,2 µmoles/litro (25 µg/dl), la
         relación entre el Pb-H y el CI puede ser diferente. Las
         estimaciones de la magnitud del efecto constituyen promedios
         grupales y sólo se aplican a cada nińo de manera probabilística.

              Los estudios epidemiológicos existentes no prueban de modo
         concluyente la existencia de un umbral. Por debajo de unos
         niveles de Pb-H de 0,48-0,72 µmoles/litro (10-15 µg/dl), los
         efectos de las variables de confusión y los límites de la
         precisión de las mediciones analíticas y psicométricas aumentan
         la incertidumbre inherente a toda estimación de un efecto. Sin
         embargo, hay algunos indicios de una asociación por debajo de
         dichos niveles.

    b)   Los estudios realizados en animales respaldan la idea de una
         relación causal entre el plomo y ciertos efectos en el sistema
         nervioso; se seńalan deficiencias cognitivas a niveles de Pb-H de
         sólo 0,53-0,72 µmoles/litro (11-15 µg/dl), deficiencias que
         pueden persistir mucho después de haber terminado la exposición
         al plomo.

    c)   Puede producirse una reducción de la velocidad de conducción
         nerviosa periférica en el ser humano con niveles de Pb-H de sólo
         1,44 µmoles/litro (30 µg/dl). Además, las funciones
         sensitivomotrices pueden verse disminuidas con niveles de Pb-H de
         sólo 1,92 µmoles/litro (40 µg/dl) aproximadamente y las funciones
         del sistema nervioso autónomo (variabilidad del intervalo R-R
         electrocardiográfico) pueden verse afectadas a un nivel promedio
         de Pb-H de aproximadamente 1,68 µmoles por litro (35 µg/dl). El
         riesgo de nefropatía plúmbica aumenta en los trabajadores que
         tienen niveles de Pb-H superiores a 2,88 µmoles/litro (60 µg/dl).
         Sin embargo, estudios recientes que utilizan indicadores más
         sensibles de la función renal sugieren la aparición de efectos
         renales a niveles más bajos de exposición al plomo.

    d)   La exposición al plomo está asociada a un pequeńo aumento de la
         presión arterial. El orden de magnitud probable es que por cada
         duplicación del nivel de Pb-H (por ejemplo, de 0,8 a
         1,6 µmoles/litro, es decir, de 16,6 a 33,3 µg/dl) hay un aumento
         medio de 1 mmHg de presión arterial sistólica. La relación con la
         presión diastólica es de una magnitud semejante pero más pequeńa.
         Sin embargo, no se sabe bien si estas asociaciones estadísticas
         obedecen realmente a un efecto de la exposición al plomo o son un
         resultado ficticio debido a factores de confusión.

    e)   Algunos estudios epidemiológicos, no todos, muestran una
         relación, dependiente de la dosis, con el parto prematuro y
         algunos índices de crecimiento y maduración fetales a niveles de
         Pb-H de 0,72 µmoles/litro (15 µg/dl) o más.

    f)   Los indicios de carcinogenicidad del plomo y de varios compuestos
         inorgánicos de plomo en el ser humano son insuficientes.

    g)   Se ha demostrado que el plomo tiene efectos en cierto número de
         sistemas enzimáticos y de parámetros bioquímicos. Los niveles de
         Pb-H por encima de los cuales las técnicas vigentes pueden
         demostrar la presencia de efectos en relación con los parámetros
         de importancia clínica posible son todos superiores a
         0,96 µmoles/litro (20 µg/dl). Algunos efectos en enzimas pueden
         demostrarse con niveles de Pb-H más bajos, pero su importancia
         clínica es incierta.
    


    See Also:
       Toxicological Abbreviations