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    UNITED NATIONS ENVIRONMENT PROGRAMME
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    WORLD HEALTH ORGANIZATION


    INTERNATIONAL PROGRAMME ON CHEMICAL SAFETY



    ENVIRONMENTAL HEALTH CRITERIA 189





    Di-n-butyl Phthalate








    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.


    Environmental Health Criteria  189


    First draft prepared by Dr G. Long and Dr E. Meek, Health and Welfare,
    Canada

    Published under the joint sponsorship of the United Nations
    Environment Programme, the International Labour Organisation, and the
    World Health Organization, and produced within the framework of the
    Inter-Organization Programme for the Sound Management of Chemicals.


    World Health Organization
    Geneva, 1997

         The International Programme on Chemical Safety (IPCS) is a joint
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    of the biological action of chemicals.

    WHO Library Cataloguing in Publication Data
     
    Di-n-butyl phthalate.

    (Environmental health criteria ; 189)

       1.Phthalic acids - adverse effects  2.Phthalic acids - toxicity
       3.Plasticizers - adverse effects    4.Plasticizers - toxicity
       5.Occupational exposure             I.Series

       ISBN 92 4 157189 6        (NLM Classification: QV 612)
       ISSN 0250-863X

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    CONTENTS

    ENVIRONMENTAL HEALTH CRITERIA FOR DI- n-BUTYL PHTHALATE

    Preamble

    1. SUMMARY

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

         2.1. Identity
         2.2. Physical and chemical properties
         2.3. Conversion factors
         2.4. Analytical methods

    3. SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE

         3.1. Natural occurrence
         3.2. Anthropogenic sources
              3.2.1. Production levels
              3.2.2. Uses
              3.2.3. Emissions

    4. ENVIRONMENTAL TRANSPORT, DISTRIBUTION AND TRANSFORMATION

         4.1. Transport and distribution between media
         4.2. Transformation
              4.2.1. Abiotic degradation
              4.2.2. Biodegradation
              4.2.3. Bioaccumulation

    5. ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE

         5.1. Environmental levels
              5.1.1. Air
              5.1.2. Water
                     5.1.2.1  Surface water
                     5.1.2.2  Groundwater
                     5.1.2.3  Seawater
                     5.1.2.4  Precipitation
                     5.1.2.5  Effluent and wastewater
              5.1.3. Sewage sludge
              5.1.4. Soil
              5.1.5. Sediment
              5.1.6. Aquatic organisms
              5.1.7. Terrestrial organisms
         5.2. General population exposure
              5.2.1. Ambient air
              5.2.2. Indoor air
              5.2.3. Drinking-water
              5.2.4. Food

              5.2.5. Consumer products
              5.2.6. Medical devices
              5.2.7. Levels in human tissue
         5.3. Occupational exposure

    6. KINETICS AND METABOLISM IN LABORATORY ANIMALS AND HUMANS

         6.1. Absorption, distribution and excretion
              6.1.1. Dermal
              6.1.2. Ingestion
                     6.1.2.1   In vivo studies
                     6.1.2.2   In vitro studies
              6.1.3. Inhalation
         6.2. Metabolic transformation
              6.2.1.  In vivo studies
              6.2.2.  In vitro studies

    7. EFFECTS ON LABORATORY MAMMALS AND  IN VITRO TEST SYSTEMS

         7.1. Single exposure
         7.2. Short-term exposure
         7.3. Long-term exposure
         7.4. Irritation and sensitization
         7.5. Reproductive and developmental toxicity
              7.5.1. Reproductive effects
                     7.5.1.1  Testicular effects
                     7.5.1.2  Effects on fertility
              7.5.2. Developmental effects
         7.6. Mutagenicity and related end-points
         7.7. Carcinogenicity
         7.8. Special studies
              7.8.1. Induction of metabolizing enzymes

    8. EFFECTS ON HUMANS

         8.1. General population exposure
         8.2. Occupational exposure
              8.2.1. Acute toxicity
              8.2.2. Epidemiological studies

    9. EFFECTS ON OTHER ORGANISMS IN THE LABORATORY AND FIELD

         9.1. Laboratory experiments
              9.1.1. Microorganisms
              9.1.2. Aquatic organisms
                     9.1.2.1  Algae
                     9.1.2.2  Invertebrates
                     9.1.2.3  Vertebrates
              9.1.3. Terrestrial organisms
                     9.1.3.1  Plants
                     9.1.3.2  Invertebrates
                     9.1.3.3  Vertebrates

    10. EVALUATION OF HUMAN HEALTH RISKS AND EFFECTS ON THE ENVIRONMENT

         10.1. Evaluation of human health risks
              10.1.1. Exposure
              10.1.2. Health effects
              10.1.3. Guidance values
         10.2. Evaluation of effects in the environment
              10.2.1. Exposure
              10.2.2. Effects
              10.2.3. Risk evaluation

    11. CONCLUSIONS AND RECOMMENDATIONS FOR PROTECTION OF HUMAN HEALTH
         AND THE ENVIRONMENT

    12. FURTHER RESEARCH

    13. PREVIOUS EVALUATIONS BY INTERNATIONAL BODIES

    REFERENCES

    RESUME

    RESUMEN
    

    NOTE TO READERS OF THE CRITERIA MONOGRAPHS

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                                     * * *

         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

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    FIGURE 1

    WHO TASK GROUP ON ENVIRONMENTAL HEALTH CRITERIA FOR DI- n-BUTYL
    PHTHALATE

     Members

    Dr B. Butterworth, Chemical Industry Institute of Toxicology Research
       Triangle Park, North Carolina, USA  (Chairman)

    Mr P. Howe, Institute of Terrestrial Ecology, Monks Wood
       Experimental Station, Abbots Ripton, Huntingdon  Cambridgeshire,
       United Kingdom  (Co-Rapporteur)

    Mr G. Long, Health and Welfare Canada, Environmental Health
       Centre, Tunney's Pasture, Ottawa, Ontario, Canada
        (Co-Rapporteur)

    Dr R. Maronpot, Laboratory of Experimental Pathology,  National
       Institute of Environmental Health Sciences, Research Triangle Park,
       North Carolina, USA

    Dr E. Meek, Health and Welfare Canada, Environmental Health Centre,
       Tunney's Pasture, Ottawa, Ontario, Canada
        (Co-Rapporteur)

    Dr S. Oishi, Department of Toxicology, Tokyo Metropolitan Research
       Laboratory of Public Health, Tokyo, Japan

    Dr Choon-Nam Ong, Department of Community, Occupational  and Family
       Medicine, National University of Singapore, Singapore

    Dr S.A. Soliman, Department of Pesticide Chemistry, Faculty of
       Agriculture, Alexandria University, El-Shatby, Alexandria, Egypt*

    Dr S.P. Srivastava, Industrial Toxicology Research Center, Lucknow,
       India

    Dr F. Sullivan, Division of Pharmacology and Toxicology, St. Thomas's
       Hospital, London, United Kingdom

    Dr C. Weber, Federal Environmental Agency, Berlin, Germany

     Secretariat

    Dr B.H. Chen, International Programme on Chemical Safety, World Health
       Organization, Geneva, Switzerland  (Secretary)

              

    *Invited but unable to attend

    ENVIRONMENTAL HEALTH CRITERIA FOR DI- n-BUTYL PHTHALATE

         A WHO Task Group on Environmental Health Criteria for
    Di- n-butyl Phthalate (DBP) met in Geneva from 30 October to
    3 November 1995.  Dr B.H. Chen, IPCS, opened the meeting and welcomed
    the participants on behalf of the Director, IPCS, and the three IPCS
    cooperating organizations (UNEP/ILO/WHO).  The Task Group reviewed and
    revised the draft criteria monograph and made an evaluation of the
    risks for human health and the environment from exposure to DBP.

         The first draft of this monograph was prepared by Dr G. Long and
    Dr E. Meek, Health and Welfare, Canada.  The second draft was prepared
    by Dr E. Meek incorporating comments received following the
    circulation of the first draft to the IPCS Contact Points for
    Environmental Health Criteria monographs.  Dr E. Meek, Mr P. Howe and
    Dr F. Sullivan contributed to the final text of this monograph.

         Dr B.H. Chen and Dr P.G. Jenkins, both members of the IPCS
    Central Unit, were responsible for the overall scientific content and
    technical editing, respectively.

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

    ABBREVIATIONS

    AP        alkaline phosphatase
    DBP       di- n-butyl phthalate
    DEHP      diethylhexyl phthalate
    GOT       glutamic-oxaloacetic transaminase
    GPT       glutamic-pyruvic transaminase
    LOAEL     lowest-observed-adverse-effect level
    LOEL      lowest-observed-effect level
    MBP       monobutyl phthalate
    NOAEL     no-observed-adverse-effect level
    NOEL      no-observed-effect level

    1.  SUMMARY

         Di- n-butyl phthalate (DBP) is an inert, colourless, oily
    liquid, with a low vapour pressure, which is soluble in most organic
    solvents, but only slightly soluble in water.  The most sensitive and
    selective analytical determinations of phthalic acid esters, including
    DBP, in environmental media are achieved by gas chromatography with
    electron capture detection or mass spectrometry.  Since phthalates
    frequently occur as plasticizers in analytical equipment and as
    contaminants in laboratory air and solvents, a great deal of care is
    needed to prevent contamination during the collection, storage and
    analysis of samples.

         DBP is used mainly as a speciality plasticizer for nitro-
    cellulose, polyvinyl acetate and polyvinyl chloride, a lubricant for
    aerosol valves, an antifoaming agent, a skin emollient and a
    plasticizer in nail polish, fingernail elongators and hair spray.

         In the atmosphere, DBP has been measured in both the vapour and
    the particulate phases. Washout via rainfall or dry deposition is
    believed to play a significant role in the removal of DBP from the
    atmosphere.  In surface water, most of the DBP is present in the water
    fraction rather than in the suspended solids.  Volatilization of DBP
    from soil is not expected to be significant because of its low vapour
    pressure and moderate adsorption to soil.

         DBP is relatively non-persistent in air and surface waters, and
    has a half-life in these compartments of only a few days.  Complete
    biodegradation of DBP is rapid under aerobic conditions but much
    slower under anaerobic conditions.  For soil, similar half-lives to
    air and water have been predicted; however, some studies suggest that
    DBP may be more persistent in soil. DBP would be expected to
    bioaccumulate as a result of its high octanol-water partition
    coefficient.  However, it is quite readily metabolized in fish and,
    consequently, bioconcentration factors tend to be lower then
    predicted.  The highest bioconcentration factor, based on the parent
    compound (DBP), is 590 for the fathead minnow.  Biomagnification is
    unlikely in terrestrial animals, based upon limited data on birds and
    the rapid metabolism and excretion observed in laboratory mammals.

         Steps taken to avoid contamination are rarely described in
    reports of concentrations of DBP in the environment published before
    1980 and, consequently, the reliability of the early monitoring data
    often cannot be assessed.  Limited data on concentrations in ambient
    air indicate that mean levels are generally less than 5 ng/m3.  In
    recent studies, mean rainwater concentrations ranged from 0.2 to
    1.4 µg/litre; much lower values have been measured in remote
    areas.  Mean concentrations in surface water tend to be less than
    1 µg/litre; however, levels in polluted rivers are much higher (12 to
    34 µg/litre).  There are only a few data on groundwater concentrations
    of DBP, mean values being 0.15 to 0.46 µg/litre.  DBP concentrations
    in effluents range up to 100 µg/litre, whilst concentrations in sewage

    sludge range from 0.2 to 200 mg/kg dry weight.  Levels in sediment are
    generally less than 1 mg/kg dry weight; however, in polluted areas
    concentrations of up to 20 mg/kg have been measured.  In studies on
    aquatic biota, mean concentrations of DBP tend to be less than
    0.2 mg/kg wet weight; however, in polluted areas, concentrations of up
    to 35 mg/kg have been measured.

         In a survey of 125 homes in California, USA, in 1990, the median
    daytime concentration of DBP in indoor air was 420 ng/m3.  DBP has
    rarely been detected in drinking-water supplies (< 1.0 µg/litre),
    according to limited data from Canada.  In a small number of samples
    of drinking-water in Toronto, Canada, the mean concentration was
    14 ng/litre; concentrations in seven brands of bottled spring water
    ranged from 21 to 55 ng/litre.

         In addition to entry through environmental contamination, DBP may
    be present in foodstuffs as a result of migration from packaging, and
    this was investigated in a number of studies conducted in the late
    1980s.  In many countries, precautions were introduced to reduce
    leaching of plasticizers from packaging and as a result, levels of DBP
    in foodstuffs have declined over time.  In a Canadian market-basket
    survey of 98 different food type  samples in Halifax in 1986, DBP was
    detected in butter (1.5 µg/g), freshwater fish (0.5 µg/g), cereal
    products (range from undetectable to 0.62 µg/g), baked potatoes
    (0.63 µg/g), coleslaw (0.11 µg/g), bananas (0.12 µg/g), blueberries
    (0.09 µg/g), pineapples (0.05 µg/g), margarine (0.64 µg/g), white
    sugar (0.2 µg/g) and gelatin dessert (0.09 µg/g).

         On the basis of the limited data available, the principal media
    of exposure to DBP for the general population, listed in order of
    their relative importance based upon estimated intake, are as follows: 
    food, indoor air and drinking-water.  Estimated intakes from food and
    indoor air are 7 µg/kg body weight per day and 0.42 µg/kg body weight
    per day, respectively.  Estimated intakes from drinking-water and
    ambient air are considerably less, < 0.02 µg/kg body weight per day
    and 0.26-0.36 ng/kg body weight per day, respectively.  Based on these
    intakes, it is estimated that the total average daily intake from air,
    drinking-water and food is 7.4 µg/kg body weight per day.  It
    should be noted, however, that intake of DBP in the diet can vary
    considerably, depending upon the nature and extent of packaged food
    consumed and the nature of use of food wrapping in food preparation. 
    For the United Kingdom, the maximum likely human intake of DBP from
    food sources has been estimated to be approximately 2 mg per person
    per day (approximately 31 µg/kg body weight per day, assuming a mean
    body weight of 64 kg).  There is also potential for exposure to DBP in
    cosmetics, although available data are inadequate to quantify intake
    from this source.

         The most recent provisional data from the NIOSH National
    Occupational Exposure Survey indicates that in the USA over 500 000
    workers, including 200 000 women, are potentially exposed to DBP.
    Based on determinations at a limited number of worksites in the USA,
    concentrations are  generally less than the limit of detection (i.e.,

    0.01 to 0.02 mg/m3), although higher levels have been reported in
    some countries.

         In studies on rats, DBP is absorbed through the skin, although in
     in vitro studies human skin has been found to be less permeable than
    rat skin to this compound. Studies in laboratory animals indicate that
    DBP is rapidly absorbed from the gastrointestinal tract, distributed
    primarily to the liver and kidneys of rats and excreted in urine as
    metabolites following oral or intravenous administration.  Following
    inhalation, it was consistently detected at low concentrations in the
    brain.

         Available data indicate that in rats, following ingestion, DBP is
    metabolized by nonspecific esterases mainly in the small intestine
    to yield mono- n-butyl phthalate (MBP) with limited subsequent
    biochemical oxidation of the alkyl side chain of MBP.  MBP is stable
    and resistant to hydrolysis of the second ester group.  The MBP and
    other metabolites are excreted in the urine mainly as glucuronide
    conjugates.  Species differences in the excretion of conjugates and
    unconjugated metabolites of DBP in the urine of rats and hamsters have
    been observed, with more free MBP being present in rats than hamsters. 
    Accumulation has not been observed in any organ.

         The profile of effects following exposure to DBP is similar to
    that of other phthalate esters, which, in susceptible species, can
    induce hepatomegaly, increased numbers of hepatic peroxisomes, 
    fetotoxicity, teratogenicity and testicular damage.

         The acute toxicity of DBP in rats and mice is low.  Reported
    LD50 values following oral administration to rats range from
    approximately 8 g/kg body weight to at least 20 g/kg body weight; in
    mice, values are approximately 5 g/kg body weight to 16 g/kg body
    weight.  The dermal LD50 in rabbits is > 4 g/kg body weight. 
    Reports of acute toxicity following inhalation of DBP have not been
    identified.  Signs of acute toxicity in laboratory animals include
    depression of activity, laboured breathing and lack of coordination. 
    In a case of accidental poisoning of a worker who ingested
    approximately 10 grams of DBP, recovery was gradual within two weeks
    and complete after 1 month.

         In short-term repeated-dose toxicity studies, effects at lowest
    levels in rats after oral administration for 5 to 21 days included
    peroxisome proliferation and hepatomegaly at doses of 420 mg/kg body
    weight per day or more.

         In longer-term studies, the effects in rats observed following
    ingestion of DBP for periods up to 7 months included reduced rate of
    weight gain at doses of 250 mg/kg body weight per day or more. 
    Increase in relative liver weight has been observed at doses of
    120 mg/kg body weight or more.  Peroxisomal proliferation with
    increased peroxisomal enzyme activity has been observed at doses of
    279 mg/kg body weight per day or more. Necrotic hepatic changes in

    Wistar rats have been reported at doses of 250 mg/kg body weight per
    day or more but not in F-344 or Sprague-Dawley rats exposed to up to
    2500 mg/kg body weight per day.  Alteration in testicular enzymes and
    degeneration of testicular germinal cells of rats have been observed
    at doses of 250 and 571 mg/kg body weight per day. There are
    considerable species differences in effects on the testes following
    exposure to DBP, minimal effects being observed in mice and hamsters
    at doses as high as 2000 mg/kg body weight per day.  In mice, effects
    on body and organ weights and histological alterations in the liver
    indicative of metabolic stress have been reported in a recent
    subchronic bioassay, for which the no-observed-effect-level (NOEL) was
    353 mg/kg body weight per day.

         On the basis of limited available data in animal species, DBP
    appears to have little potential to irritate skin or eyes or to induce
    sensitization.  In humans, a few cases of sensitization after exposure
    to DBP have been reported, although this was not confirmed in
    controlled studies of larger numbers of individuals reported only in
    secondary accounts.

         In a continuous breeding protocol, which included cross-over
    mating and offspring assessment phases, rats were exposed to 0, 1000,
    5000 or 10 000 mg DBP/kg in the diet (equivalent to 0, 66, 320 and
    651 mg/kg body weight per day).  In the first generation the reduction
    in pup weight in the mid-dose group, in the absence of any adverse
    effect on maternal weight, could be regarded as a developmental
    toxicity effect.  There was also a significant reduction of live
    litter numbers at all three dose levels.  The effects in the second
    generation were more severe, with reduced pup weight in all groups
    including the low-dose group, structural defects (such as prepucial/
    penile malformations, seminiferous tubule degeneration, and absence or
    underdevelopment of the epididymides) in the mid- and high-dose
    groups, and severe effects on spermatogenesis in the high-dose group
    that were not seen in the parent animals.  These results suggest that
    the adverse effects of DBP are more marked in animals exposed during
    development and maturation than in animals exposed as adults only.  No
    clear NOEL was established in this study.  The lowest-observed-
    adverse-effect-level (LOAEL) was considered to be 66 mg/kg body weight
    per day.

         The available studies show that DBP generally induces fetotoxic
    effects in the absence of maternal toxicity.  Available data also
    indicate that DBP is teratogenic at high doses and that susceptibility
    to teratogenesis varies with developmental stage and period of
    administration. In mice, DBP caused dose-dependent increases in the
    number of resorptions and dead fetuses at oral doses of 400 mg/kg body
    weight per day or more. Dose-dependent decreases in fetal weights and
    number of viable litters were also observed in mice at these doses.

         Adequate carcinogenesis bioassays for DBP have not been
    conducted.  The weight of the available evidence indicates that DBP is
    not genotoxic.

         As a class, chemicals which cause peroxisome proliferation are
    often hepatocarcinogenic via a non-genotoxic mode of action.  Although
    the mechanism of action remains unknown, tumour formation is preceded
    by peroxisomal proliferation and hepatomegaly.  Since DBP causes
    peroxisomal proliferation, it is possible that it might be a rodent
    liver carcinogen, although it is much weaker in inducing hepatomegaly
    and peroxisome proliferation than DEHP.  To the degree that
    hepatomegaly and peroxisomal proliferation correlate with carcinogenic
    potency, DBP would be expected to be a less potent carcinogen than
    DEHP and would probably exhibit no activity as measured by current
    cancer bioassay methodologies.

         Identified epidemiological investigations are limited to those of
    workers exposed to mixtures of phthalates.  These studies do not
    contribute to our understanding of the effects associated with DBP
    alone.

         Since DBP is not genotoxic and is expected to be a less potent
    carcinogen than DEHP, it would probably exhibit no activity as
    measured by current cancer bioassay methodologies.  Thus, it is
    unlikely that DBP presents any significantly increased risk of cancer
    at concentrations generally present in the environment.

         Ingestion is by far the principal route of exposure to DBP;
    moreover, the toxicological data for other routes of administration
    are insufficient for evaluation. A guidance value has, therefore, been
    developed for the oral route, although the ultimate objective should
    be reduction of total exposure from all sources to less than the
    tolerable daily intake.

         No clear no-observed-adverse-effect-level (NOAEL) for the
    end-points considered to be most appropriate for derivation of
    guidance values (i.e., developmental and reproductive toxicity) was
    established.  The LOAEL for developmental and reproductive toxicity
    from a continuous breeding study was considered to be 66 mg/kg body
    weight per day, although the effects observed at this dose level were
    moderate and probably reversible.  On the basis of these data, a
    tolerable daily intake of 66 œg/kg body weight per day has been
    derived, incorporating an uncertainty factor of 1000 (× 10 for
    interspecies variation, × 10 for inter-individual variation, and × 10
    for extrapolation from LOAEL to NOAEL).

         Information on the ecotoxicity of DBP includes acute and chronic
    data for a number of species from various trophic levels in the
    aquatic environment.  For freshwater algae the lowest reported 96-h
    EC50 was 750 µg DBP/litre.  The lowest reported values in acute
    toxicity tests on aquatic invertebrates were a 96-h LC50 of
    750 µg/litre (mysid shrimp) and a 48-h EC50 of 760 µg/litre (midge
    larvae).  In  chronic studies, the most sensitive invertebrate species
    was  Daphnia magna, with a 21-day NOEC (parent survival) of
    500 µg/litre.  In a non-standard test with the scud  (Gammarus pulex)
    a 10-day LOEC of 500 µg/litre and a NOEC of 100 µg/litre, both based 

    on reduced locomotor activity, were reported.  In acute toxicity tests
    with fish the lowest reported 96-h LC50 for a freshwater species was
    350 µg/litre (yellow perch) and for a marine species 600 µg/litre
    (sheepshead minnow).  The most sensitive chronic study was based on
    the rainbow trout with a 99-day NOEC (growth) of 100 µg/litre and a
    99-day LOEC of 190 µg/litre (growth reduced by about 27%).

         The acute toxicity of DBP to birds is low.

         The risk to aquatic organisms associated with the present mean
    concentrations of DBP in surface water is low.  However, in highly
    polluted rivers the safety margin is much smaller.  There is
    inadequate data to assess the risk of DBP to sediment-dwelling
    organisms.  At current levels of exposure, it can be concluded that
    the risk to fish-eating birds and mammals is low.

    2.  IDENTITY, PHYSICAL AND CHEMICAL PROPERTIES AND ANALYTICAL METHODS

    2.1  Identity

         Di- n-butyl phthalate (DBP), a phthalic acid ester, has the CAS
    (Chemical Abstracts Service) Registry Number 84-74-2, the molecular
    formula C16H22O4, and a relative molecular mass of 278.4.  Synonyms
    and trade names are presented in Table 1.

    2.2  Physical and chemical properties

         DBP is an inert colourless oily liquid, with a vapour pressure of
    about 0.01 Pa at 25°C (CMA, 1984), Henry's law constant of 4.6 × 10-7
    atmÊm3/mol at 25°C (Howard, 1989) and an octanol-water partition
    coefficient (log Kow) between 4.31 and 4.79 (Montgomery & Welkom,
    1990). The solubility in water is about 10 mg/litre (McKone & Layton,
    1986), although higher values have also been reported (Montgomery &
    Welkom, 1990).  The determination of the water solubility of phthalic
    acid esters is complicated since these compounds easily form colloidal
    dispersions (Klöpfer et al., 1982) and are subject to "molecular
    folding" (Callahan et al., 1979).  DBP is soluble in most of the
    organic solvents (BUA, 1987).  Additional chemical and physical
    properties of DBP are presented in Table 1.

    2.3  Conversion factors

    1 ppm = 11.4 mg/m3
    1 mg/m3 = 0.088 ppm

    2.4  Analytical methods

         The most sensitive and selective analytical determinations of
    phthalic acid esters, including DBP, in environmental media are
    achieved by gas chromatography (GC) with electron-capture detection
    (ECD), with or without derivatization (Kohli et al., 1989).  In the
    analysis of environmental samples it is imperative to note that peaks
    of other components can interfere with determinations of DBP.  This
    problem is particularly serious when ECD is used, because of its high
    sensitivity towards halogenated aromatics, PCBs etc.  The US
    Environmental Protection Agency has standardized sample preparation
    and analysis for municipal and industrial wastewater using GC with ECD
    (Method 606, detection limit 0.36 µg/litre) and GC/mass spectrometry
    (MS) (Method 625, detection limit 2.5 µg/litre) (US EPA, 1982b). 
    Thin-layer chromatography may be used to separate phthalates from
    other solvent-extracted organic compounds.  Analysis can also be
    carried out by using high-performance liquid chromatography with
    ultraviolet detection (HPLC-UV) (Poole & Wibberley, 1977).

    Table 1.  Physical properties of di- n-butyl phthalate
              (Adapted and modified from: USEPA, 1981; ATSDR, 1990)
                                                                        

    Chemical formula         C16H22O4

    Structure

    CHEMICAL STRUCTURE 1

    Relative molecular mass  278.34

    Synonyms                 butylphthalate; dibutylphthalate; DBP;
                             1,2-benzenedicarboxylic acid dibutyl ester;
                              o-benzenedicarboxylic acid, dibutyl ester;
                             dibutyl 1,2-benzene dicarboxylate;
                             dibutyl- o-phthalate

    CAS name                 1,2-benzenedicarboxylic acid, dibutyl ester

    CAS registry number      84-74-2

    Trade names              Caswell No. 292; Uniflex DBP; Celluflex DBP;
                             Ergoplast FDB; Polycizer DBP; Genoplast B;
                             Staflex DBP; Palatinol C; Hexaplast M/B; PX
                             104; RC Plasticizer DBP

    Physical state           Oily liquid

    Colour                   Colourless

    Odour                    Mild, aromatic

    Melting point            -35°C

    Boiling point            340°C

    Flashpoint               171°C
                                                                        

    Table 1.  contd. 

                                                                        

    Vapour pressure at 25°C  0.01 Pa (1.0 × 10-5 mmHg)

    Density at 20°C          1.047

    Partition coefficients
      Log octanol/water      4.31-4.79
      Log Koc                5.23

    Solubility
      Water at 25°C          10 mg/litre
      Organic solvents       Soluble in alcohol, ether, benzene

    Henry's law constant     4.6 × 10-7 atmÊm3/mol
                                                                        

         Phthalates frequently occur as plasticizers in analytical
    equipment and as contaminants in laboratory air and solvents.  This
    can result in overestimation of their concentration in environmental
    samples.  For example, Ishida et al. (1980) detected DBP in laboratory
    solvents at concentrations as high as 0.17 mg/kg (in benzene)
    and in solid reagents at concentrations up to 9.89 mg/kg (in
    carboxymethylcellulose), while polyvinyl tubing contained 20% DBP. 
    Therefore, a great deal of care is needed to prevent contamination
    during the collection, storage and analysis of samples (Mathur, 1974;
    US EPA, 1982b; Kohli et al., 1989; Hites & Budde, 1991).  A summary of
    analytical methods for the determination of DBP in environmental
    samples and biological materials is presented in Tables 2 and 3,
    respectively.

        Table 2.  Analytical methods for determining di- n-butyl phthalate in environmental samplesa

                                                                                                                                 

    Sample matrix              Sample preparation          Analytical   Sample detection
                                                            methodsb          limit        Accuracy         Reference
                                                                                                                                 

    Air                   Adsorption/solvent extraction      HRGC/MS      No data          115 ± 5%c     Ligocki & Pankow
                          with polyurethane foam plug                                                    (1985)

    Rainwater             Adsorb on Tenax-GC columns,        GC/MS        < 34 ng/litre    No data       Ligocki et al.
                          thermally desorb                                                               (1985)

    Water                 Extract with dichloromethane,      GC/ECD       0.36 µg/litre    80 ± 6%c      US EPA (1982a)
                          exchange to hexane, concentrate

    Water                 Extract with dichloromethane at    GC/MS        2.5 µg/litre     80 ± 6%c      US EPA (1982b)
                          pH 11 and 2, concentrate

    Water                 Adsorb on small bed volume         GC/MS        No data          No data       Pankow et al.
                          Tenax cartridges, thermally                                                    (1988)
                          desorb

    Soil                  Extract with dichloromethane,      GC/ECD       240 ng/kg        96%           US EPA (1986a)
                          clean up, exchange to hexane

    Waste,                Extract with dichloromethane,      GC/ECD       36 mg/kg         96%           US EPA (1986a)
    non-water-miscible    clean up, exchange to hexane

    Soil                  Extract from sample, clean up      GC/MS        1.7 mg/kg        96%           US EPA (1986b)

    Waste,                Extract from sample, clean up      GC/MS        350 mg/kg        76%           US EPA (1986b)
    non-water-miscible

    Soil/sediment         Extract from sample, clean up      HRGC/MS      660 µg/kg        76%           US EPA (1986c)
                                                                                                                                 

    Table 2.  Continued

                                                                                                                                 

    Sample matrix              Sample preparation          Analytical   Sample detection
                                                            methodsb          limit        Accuracy         Reference
                                                                                                                                 

    Waste,                Extract from sample, clean up      HRGC/MS      50 mg/kg         76%           US EPA (1986c)
    non-water-miscible

    Soil/sediment         Extract from sample, clean up      HRGC/FTIR    10 µg/litred     No data       US EPA (1986d)

    Wastes,               Extract from sample, cleanup       HRGC/FTIR    10 µg/litred     No data       US EPA (1986d)
    non-water-miscible
                                                                                                                                 

    a    From: Agency for Toxic Substances and Diseases Registry (1990).
    b    HRGC =  high-resolution gas chromatography;
         MS   =  mass spectrometry;
         GC   =  gas chromatography;
         ECD  =  electron-capture detector;
         FTIR =  Fourier transform infrared spectrometry.
    c    Relative recovery, percentage ± standard deviation.
    d    Identification limit.  Detection limits for actual samples are several orders of magnitude higher depending upon the sample
         matrix and extraction procedure employed.

    Table 3.  Analytical methods for determining di- n-butyl phthalate in biological materials

                                                                                                                               

    Sample matrix           Sample preparation       Analytical     Sample detection    Accuracy         Reference
                                                       methoda            limit       (% recovery)
                                                                                                                               

    Aquatic organisms    Extract with acetonitrile    HRGC/ECD         0.1 µg/kg         68           Thuren (1986)
                         and petroleum ether

    Adipose tissue       Extraction, bulk lipid       HRGC/MS          10 µg/kg          No data      Stanley (1986)
                         removal, Florisil
                         fractionation

    Blood serum          Extraction, bulk lipid       HRGC/MS          10 µg/kg          No data      Stanley (1986)
                         removal, Florisil
                         fractionation

    Blood serum          Extraction with organic      GC/MS            No data           No data      Ching et al. (1981)
                         solvents (propanol,
                         heptane)

    Cooked meat          Remove with nitrogen gas     GC/MS            No data           No data      Ho (1983)
                         trap, extract with diethyl
                         ether
                                                                                                                               

    a    HRGC   High-resolution gas chromatography;
         ECD    Electron-capture detector;
         MS     Mass spectrometry;
         GC     Gas chromatography
    
    3.  SOURCES OF HUMAN AND ENVIRONMENTAL EXPOSURE

    3.1  Natural occurrence

         The occurrence of naturally produced phthalates in biological and
    geochemical samples has been suggested, but in most cases the
    possibility of contamination during sampling or analysis could not be
    ruled out (Mathur, 1974).  However, it is unlikely that the amounts of
    phthalates produced naturally would be significant compared with those
    from anthropogenic sources (IPCS, 1992).

    3.2  Anthropogenic sources

    3.2.1  Production levels

         Total DBP production in western Europe in 1994 was estimated to
    be 49 000 tonnes (personal communication by the European Council for
    Plasticisers and Intermediates to the IPCS, 1996).  In Germany, the
    average annual production was 20 000 tonnes for 1982-1986 (BUA, 1987). 
    DBP is produced by 36 companies in the USA, with total production of
    7720 tonnes in 1977 and 11 400 tonnes in 1987 (ATSDR, 1990; NTP,
    1995).  Annual production in Japan in 1994 was  about 17 000 tonnes
    (JPIF, 1995).

    3.2.2  Uses

         DBP is used mainly as a speciality plasticizer for nitrocellulose
    polyvinyl acetate and polyvinyl chloride (PVC) (ATSDR, 1990).  In
    1991, approximately 54% of the total supply of DBP in Canada was used
    in adhesives, while about 15% was used in coatings (including
    lacquers), and the rest in miscellaneous applications, including paper
    coating (Camford Information Services Inc., 1992).

         In Germany, approximately 25% of the DBP produced served as
    plasticizer and adjuvant for the processing of PVC and about 20% was
    used in adhesives (BUA, 1987).

         DBP is one of the most commonly used plasticizers in regenerated
    cellulose film, being present mainly in nitrocellulose coatings which
    are applied to the films (average content, 2.5% of the  weight of the
    film) (MAFF, 1987).

         DBP is used in cosmetics as a perfume solvent and fixative, a
    suspension agent for solids in aerosols, a lubricant for aerosol
    valves, an antifoaming agent, a skin emollient and a plasticizer in
    nail polish, fingernail elongators and hair spray (Brandt, 1985).

    3.2.3  Emissions

         Although DBP has low volatility, its widespread use in many thin
    polymeric sheets and coatings provides large surface areas for
    volatization during manufacture, use and disposal of these products.

    Disposal at dump sites and disintegration or incineration of the
    plastics allow for dispersal of small particulates into the air
    (ATSDR, 1990)  Perwak et al. (1981) estimated that about 300 tonnes of
    DBP were released into the air in 1977 in the USA.

         Based on a production of 22 100 tonnes in Germany in 1986,
    the release into the environment was estimated to be about
    500 tonnes/year.  Release associated with the production of DBP was
    estimated to be about 0.1 tonnes/year, whereas emission related
    to end usage was 400 tonnes/year.  It was estimated that about
    100 tonnes/year were released by further processing activities, such
    as manufacture of plastic and other materials (BUA, 1987).

         DBP may be released into surface water.  It is estimated that
    300 tonnes of DBP were released to water in 1977 in the USA (Perwak et
    al., 1981).

         No specific release of DBP to soils has been reported.  However,
    it may seep into soil from DBP coating sewage sludge that is deposited
    on land (ATSDR, 1990).

    4.  ENVIRONMENTAL TRANSPORT, DISTRIBUTION AND TRANSFORMATION

    4.1  Transport and distribution between media

         In the atmosphere, DBP has been measured in both the vapour and
    the particulate phases.  In various studies, the proportion of total
    DBP present in the vapour form in the atmosphere has been reported to
    range from 68% (32% in the particulate phase) in the Gulf of Mexico
    (Giam et al., 1980) to 78% (22% in the particulate phase) in Antwerp,
    Belgium (Cautreels & van Cauwenberghe, 1978).  Hoff & Chan (1987),
    however, reported that in the Niagara River region of North America,
    more than 57% of atmospheric DBP occurs in the suspended particulate
    phase.

         Washout via rainfall or dry deposition is believed to play a
    significant role in the removal of DBP from the atmosphere. 
    Eisenreich et al. (1981) predicted that atmospheric deposition is a
    significant source of DBP in the Great Lakes, North America, with a
    calculated total deposition of 48 tonnes/year to the five Great Lakes
    and values for each ranging from 3.7 tonnes/year for Lake Ontario to
    16 tonnes/year for Lake Superior.  Based on levels of DBP in airborne
    fallout at 14 locations in Sweden, the total deposition was estimated
    to be 90 tonnes per year (Thurén & Larsson, 1990).

         In surface water, most of the DBP (> 75%) is present in the
    water fraction rather than in the suspended solids (Niagara River Data
    Interpretation Group, 1990).  Sullivan et al. (1982) reported that DBP
    was rapidly adsorbed onto and desorbed from three clay minerals,
    sediment and glass test tubes.  During the experiments no more than
    11% of the total DBP was adsorbed.  Al-Omran & Preston (1987) found
    that DBP reached an adsorption equilibria within 30 min, the degree of
    adsorption being most closely correlated to the lipid content of
    suspended particles.  The adsorption was enhanced by the presence of
    salt.

         DBP is moderately adsorbed to soil (Howard, 1989; Zurmühl et al.,
    1991), but it forms a complex with water-soluble fulvic acid and this
    may increase its mobilization and reactivity in soil to some degree
    (Matsuda & Schnitzer, 1971).  Volatilization of DBP from soil is not
    expected to be significant because of its low vapour pressure and
    moderate adsorption to soil (Howard, 1989).

         Using the Exposure Analysis Modelling System (EXAMS), Wolfe et
    al. (1980) calculated that at equilibrium the loss of DBP from a pond
    was 3.3% hydrolysis, 1.2% photolysis, 31.8% biodegradation and 6.2%
    volatilization.

    4.2  Transformation

    4.2.1  Abiotic degradation

         Howard et al. (1991) estimated the photo-oxidation half-life of
    DBP in air to range from 7.4 h to 3.1 days.

         The photolytic half-life of DBP in water has been estimated to be
    144 days (Howard, 1989; calculated from Wolfe et al., 1980).

    4.2.2  Biodegradation

         DBP is biodegradable in natural surface waters, with an estimated
    half-life in the range of 1 to 14 days (Schouten et al., 1979; Johnson
    et al., 1984; Walker et al., 1984; Howard, 1989; Howard et al., 1991).

         Primary degradation exceeded 95% in 24 h in the Semi-Continuous
    Activated Sludge (SCAS) test, while ultimate biodegradation to CO2
    amounted to 57.4% (half-life of 15.4 days) in the shake flask test
    (CMA, 1984).  Sugatt et al. (1984) reported 90% primary degradation of
    DBP in the 28-day shake flask test using mixed populations of
    microorganisms from natural sources.

         Howard et al. (1991) predicted a DBP half-life of 2-23 days in
    groundwater, based upon aerobic and anaerobic degradation rates.

         Sediment from the upper 5 cm of a test pond served as the
    inoculum in tests of aerobic and anaerobic degradation of DBP (Johnson
    & Lulves, 1975). The samples contained 1 mg/litre of 14C-labelled
    DBP. The extent of aerobic degradation was 53% within 24 h and 98%
    within 5 days. The anaerobic solutions still contained 69% of the
    initial amount after 5 days and only 2% after 30 days.

         O'Connor et al. (1989) found > 85% mineralization of DBP during
    incubation of anaerobic sludge for 90 days at a concentration of
    200 mg DBP/litre.  In anaerobic sludge, degradation of DBP proceeded
    through mono- n-butyl phthalate to phthalic acid, followed by ring
    cleavage and mineralization (Shelton et al., 1984).

         In an experiment with batch anaerobic digestion of sewage sludge
    spiked with DBP at a concentration range of 0.5-10 mg/litre, DBP was
    degraded rapidly with a degradation rate following first-order
    kinetics.  More than 90% was removed in under 8 days without any lag
    phase (Ziogou et al., 1989).  The degradation rate can vary with
    sludge source and sampling time.  DBP was found to be degraded from an
    activated sludge system very efficiently (Iturbe et al., 1991).

         In a series of studies, Kurane et al. (1979a,b) demonstrated that
    DBP is efficiently removed from wastewater by inoculating viable cells
    of  Nocardia erythropolis, a bacterium capable of rapidly degrading
    phthalate esters in activated sludge.  When the wastewater containing
    3000 mg DBP/litre was treated with the activated sludge inoculated
    with  N. erythropolis, the DBP was found to be removed at a rate of
    94.2% in one day and 100% after the 5th day (as measured by gas

    chromatography) (Kurane et al., 1979a,b).  Phthalate ester-utilizing
    microoganism species isolated from the inoculated and uninoculated
    activated sludge were  N. erythropolis, N. restricta, Pseudomonas
     capacia,  P. fluorescens and  P. acidovorans (Kurane et al.,
    1979a,b).

          Pseudomonas pseudoalcaligenes B20b1 (a denitrifying strain) was
    enriched from the effluent of a biological sewage plant with DBP as
    the sole carbon source (Benckiser & Ottow, 1982).  After 20 days
    at 30°C, TLC and MS analysis of the culture extracts showed
    mono- n-butyl phthalate and  phthalic acid as the only products,
    suggesting that an  n-butanol moiety served essentially as the carbon
    source for growth and denitrification.  A  Micrococcus sp. (strain
    12B) was also isolated by enriching with DBP as sole carbon and energy
    source, and a metabolic pathway for DBP by this strain was proposed
    (Eaton & Ribbons 1982).  In this  pathway, DBP is converted to mono-
     n-butyl phthalate and then to 3,4-dihydro-3,4-dihydroxy phthalate,
    which is in turn converted to 3,4-dihydroxy phthalate and then to
    protocatechuate (3,4-dihydroxy benzoate).  Protocatechuate is
    metabolized by a meta-cleavage pathway to pyruvate and oxaloacetate
    and by an ortho-cleavage pathway to beta-keto-adipate (Eaton &
    Ribbons, 1982).

         Wang et al. (1995) isolated five strains of DBP-degrading
    microorganisms from coke-plant wastewater treatment plant sludge.
    All strains were capable of achieving complete degradation of DBP
    (100 mg/litre).  One strain was able to completely degrade DBP within
    40 h.  Further experimental studies revealed that the rate of DBP
    degradation was higher with immobilized cells than with free cells.

         Chauret et al. (1995) have isolated a psychrotrophic denitrifying
     Pseudomonas fluorescens from DBP-spiked microcosms, which is
    capable of transforming DBP at 10°C under both aerobic and anaerobic
    conditions.  The isolated pseudomonad did not grow with phthalic acid
    as the sole source of carbon, indicating that DBP was not mineralized
    by this bacterium.

         Howard et al. (1991) predicted a half-life for DBP in soil of 2
    to 23 days.  Inman et al. (1984) reported that DBP was almost
    completely metabolized within 100 days in non-sterile soils of various
    types (silt loam, sand, mixture of silica sand and peaty muck). 
    Overcash et al. (1982), however, reported half-lives of > 26 weeks in
    loam and sand at application rates of 800 mg DBP/kg or more, while, at
    a lower application rate (200 mg/kg), the half-life of DBP in loam and
    sand was about 12 weeks.

         Shanker et al. (1985) incubated garden soil containing DBP at a
    concentration of 500 mg/kg.  Within 10 days, 91% of the DBP had been
    degraded and, after 15 days, 100% of the parent compound had been
    degraded.  No degradation was detected when sterilized soil was used. 
    Degradation of DBP was much slower in anaerobic soil, flooded with
    sterile water to reduce oxygen tension.  After a 30-day incubation,

    66% of the DBP had been degraded, compared with 100% degradation
    within 15 days under aerobic conditions.

         Yan et al. (1995) reported that algae are capable of degrading
    DBP.  An average biodegradation rate of 2.1 mg/litre per day was found
    when the alga  Chlorella pyrenoidosa was exposed to 7 mg DBP/litre. 
    Degradation of the parent compound was complete within 72 h.

    4.2.3  Bioaccumulation

         The log octanol-water partition coefficient for DBP is between
    4.31 and 4.79, which indicates a potential for the chemical to
    bioaccumulate.  However, the accumulation of DBP is influenced by the
    capability of an organism to metabolize it, and several authors have
    shown the ability of fish to metabolize DBP.  Stalling et al. (1973)
    found that radioactively-labelled DBP was metabolized by microsomal
    preparations from fish (channel catfish) liver to mono- n-butyl
    phthalate (55%) and three other unidentified metabolites (42%) within
    2 h.  Only 3% of the parent compound was recovered.  All of the values
    are expressed as percentage of radioactivity.  The hepatic microsomes
    taken from male channel catfish degraded DBP 16 times more rapidly
    than diethylhexyl phthalate (DEHP).  When Wofford et al. (1981)
    exposed sheepshead minnow to 14C-DBP for 24 h, the distribution of
    metabolites was as follows: 13% diester; 28.2% monoester; 47.8%
    phthalic acid; and 11% of the radioactivity in the residue.

         Bioconcentration factors for a number of organisms are presented
    in Table 4.  A wide variety of bioconcentration factors have been
    reported reflecting not only the capability of organisms to accumulate
    DBP but also the variety of exposure concentrations and test
    conditions.  Care must be taken when interpreting data based on the
    accumulation of radioactivity because of the metabolism of the parent
    compound (DBP).  The highest bioconcentration factor quoted, based on
    the parent compound, is 590 for the fathead minnow ( Pimephales
     promelas) at an exposure concentration of 34.8 µg/litre. The
    bioconcentration factor was a mean value based on the percentage of
    DBP in the measured radioactivity over an 11-day period.  The
    percentage of DBP ranged from 50% on day 3 to 8% on day 11 (Call et
    al., 1983).

         Lokke & Bro-Rasmussen (1981) applied DBP, in a mixture that also
    contained DEHP and di-iso-butyl phthalate, at a concentration of
    2.5 µg/cm2 to the leaves of  Sinapis alba.  The residue level of
    DBP on the leaves immediately after application was 2.4 µg/cm2. 
    There was rapid elimination of DBP and after 15 days DBP levels had
    decreased to only 0.03 µg/cm2.

         Belisle et al. (1975) fed mallard ducks ( Anas platyrhynchos)
    on a diet containing 10 mg DBP/kg for a period of 5 months.  No DBP
    was detected in fat, heart, lung or breast tissue (detection limit = 
    0.1 mg/kg in a 2-g sample).  The exposure concentration was equivalent
    to a dose of 0.56 mg/kg body weight per day, assuming a body weight of

    1.1 kg/bird and a food consumption rate of  0.0619 kg dry weight per
    day (Nagy, 1987).  There appears to have been no biomagnification of
    DBP in this study.  In fact, it would seem unlikely that terrestrial
    animals will biomagnify DBP, based upon the rapid  metabolism and
    excretion observed in laboratory mammals (see Chapter 6).

        Table 4.  DBP bioconcentration (BCF) factors for various aquatic organisms

                                                                                        

    Species              Water      Duration        BCFa             Reference
                     concentration   (days)
                       (µg/litre)
                                                                                        

    Oyster                100           1        21.1b          Wofford et al.
     (Crassostrea                                               (1981)
      virginica)

    Oyster                500           1        41.6b          Wofford et al.
     (Crassostrea                                               (1981)
      virginica)

    Water flea           0.08          14        400c           Mayer & Sanders
     (Daphnia magna)                                            (1973)

    Scud                 0.10          14        1400c          Mayer & Sanders
     (Gammarus                                                  (1973)
      pseudolimnaeus)

    Scud                  100          10        140            Thurén & Woin (1991)
     (Gammarus pulex)                            (accumulated)

    Scud                  100          10        45             Thurén & Woin (1991)
     (Gammarus pulex)                            (adsorbed)

    Scud                  500          10        64             Thurén & Woin (1991)
     (Gammarus pulex)                            (accumulated)

    Scud                  500          10        8.4            Thurén & Woin (1991)
     (Gammarus pulex)                            (adsorbed)
                                                                                        

    Table 4.  Continued

                                                                                        

    Species              Water      Duration        BCFa             Reference
                     concentration   (days)
                       (µg/litre)
                                                                                        

    Brown shrimp          100           1        2.9            Wofford et al. (1981)
     (Penaeus aztecus)

    Brown shrimp          500           1        30.6           Wofford et al. (1981)
     (Penaeus aztecus)

    Midge                0.18           7        720c           Mayer & Sanders (1973)
     (Chironomus
      plumosus)

    Mayfly              0.008           7        430c           Mayer & Sanders (1973)
     (Hexagenia
      bilineata)

    Fathead minnow       4.83          11        570d           Call et al. (1983)
     (Pimephales
      promelas)

    Fathead minnow       34.8          11        590d           Call et al. (1983)
     (Pimephales
      promelas)

    Sheepshead minnow     100           1        11.7           Wofford et al. (1981)
     (Cyprinodon
      variegatus)
                                                                                        

    a    BCF based on whole-body concentrations, unless otherwise indicated
    b    BCF based on concentration in muscle
    c    Based on radioactivity
    d    Based on a mean for the % DBP in the radioactivity measured on days 1, 3 and 11
    
    5.  ENVIRONMENTAL LEVELS AND HUMAN EXPOSURE

    5.1  Environmental levels

         Identified data on concentrations of DBP in various media
    are presented in Table 5.  Data from the surveys considered to be
    most representative are addressed in the text.

         In interpreting this data, it should be noted that steps
    taken to avoid contamination are rarely described in the reports
    published before 1980 and, consequently, the reliability of the
    early data often cannot be assessed.  The more recent available
    data have therefore been emphasized.

    5.1.1  Air

         The levels of DBP in air are summarized in Table 5.

         Giam et al. (1978) reported mean concentrations of
    0.3 ng/m3 over the Gulf of Mexico (n = 8) and 1.0 ng/m3 over
    the North Atlantic Ocean (n = 5).  No other information was
    provided.

         DBP was detected in samples of air taken in 1982 (n = 5)
    along the Niagara River in Ontario, Canada, with mean
    concentrations of 1.9 ± 1.3 ng/m3 in the gas phase and
    4.0 ± 2.2 ng/m3 in the particulate phase (Hoff & Chan, 1987). 
    In 1983, mean levels were 4.5 ± 3.5 ng/m3 in 15 samples of
    the gas phase and 6.2 ± 2.6 ng/m3 in 19 samples of the
    particulate phase.  Eisenreich et al. (1981) reported that
    atmospheric concentrations of DBP in the Great Lakes area ranged
    from 0.5 to 5 ng/m3; however, no sampling or analytical
    details were given.

         DBP has been identified in ambient air in Barcelona, Spain;
    concentrations of 3.0 and 17 ng/m3 were reported in winter, and
    1.1 and 10 ng/m3 in summer for coarse (> 7.2 µm) and fine
    (> 0.5 µm) particulates, respectively (Aceves & Grimalt 1993).

         Cautreels et al. (1977) reported a range of concentrations
    of DBP from 24 to 74 ng/m3 in the suspended particulate phase of 
    the air in a residential area of Antwerp, Belgium, in contrast to
    19 to 36 ng/m3 in samples from a rural area in Bolivia.  Atlas &
    Giam (1981) reported atmospheric concentrations of DBP as high as
    18.5 ng/m3 at Pigeon Key, Florida.  Bove et al. (1978) reported
    mean concentrations of DBP ranging from 3.28 ng/m3 at Staten
    Island to 5.69 ng/m3 at Brooklyn, New York.  Weschler (1981)
    reported DBP in the Arctic aerosol at Barrow, Alaska, at a
    concentration of about 1 ng/m3.  In Japan, in 1985, DBP was
    detected in 56 out of 63 samples of ambient  air at levels
    ranging from 17 to 370 ng/m3 (detection limits, 5 to
    70 ng/m3) (Environment Agency, Japan, 1995).

    5.1.2  Water

    5.1.2.1  Surface water

         The levels of DBP in surface water are summarized in
    Table 5.  Information on concentrations of DBP in surface water
    in a national database in Canada is limited to 73 records for
    Alberta and two records for British Columbia dating from 1985 to
    1988.  Concentrations were above the detection limit for only
    eight records and reported values ranged from < 1 to 2 µg/litre
    (NAQUADAT, 1993).  For water samples collected in 1988 and 1989,
    mean concentrations of 12.2 ng/litre at Fort Erie, Ontario (all
    of 26 samples contained DBP at concentrations above the
    detection limit of 0.29 ng/litre; maximum 26.78 ng/litre) and
    15.16 ng/litre at Niagara-on-the-Lake, Ontario (all of 25 samples
    contained DBP at concentrations above the detection limit of
    0.29 ng/litre; maximum 72.93 ng/litre) were reported (Niagara
    River Data Interpretation Group, 1990).

         In Japan, for the years 1974, 1975 and 1982, levels of DBP
    in surface water ranged from 0.013 to 36 µg/litre (detected in 55
    to 93% of samples; detection limits, 0.01 to 40 µg/litre). 
    (Environment Agency, Japan, 1995).

         In 1991 and 1992; DBP concentrations were measured in
    unfiltered water samples of the River Rhine (4 locations) and six
    of its tributaries.  DBP was detected in 99% of 217 samples with
    a detection limit of 0.03 µg/litre.  The mean concentration in
    the Rhine was 0.18 µg/litre, and the maximum value was
    1.3 µg/litre.  Mean values in the tributaries were in the same
    range (LWA, 1993; Furtmann, 1994).  The concentrations in the
    particulate fraction of R. Rhine water were reported to be in the
    range of 1.2 to  7.8 mg/kg dry weight.  Schouten et al. (1979)
    reported that DBP  concentrations in rivers in the Netherlands
    ranged from < 0.1 to 2.8 µg/litre.  Other measurements of DBP
    concentrations in the Netherlands revealed a mean value of
    0.1 µg/litre in the Rhine (maximum = 1.1 µg/litre, 53 samples) in
    1991 (RIWA, 1991) and 1.0 µg/litre in the Ijssel Sea (maximum =
    6.9 µg/litre; 7 samples) in 1992 (RIWA, 1992). In both reports a
    mean value of 0.1 µg/litre was given for the River Lek.

         In 1984, DBP was detected in the Rivers Irwell (12.1 and
    33.5 œg/litre) and Etherow (32.5 and 23.5 œg/litre) in
    Manchester, United Kingdom (Fatoki & Vernon, 1990).  Both rivers
    received discharges from factories making plastic products.

    5.1.2.2  Groundwater

         At four sites in woodland areas of Germany, which are not
    directly influenced by industry or agriculture, DBP
    concentrations were measured monthly in wellwater and groundwater
    in 1988 and 1989 (Schleyer et al., 1991).  Mean concentrations
    were 0.15 to 0.46 µg/litre.

    5.1.2.3  Seawater

         The levels of DBP in seawater are summarized in Table 5.  In
    an early study, concentrations of DBP up to 0.47 µg/litre in
    water from the Gulf of Mexico were reported (Chan, 1975). 
    Reported maximum concentrations of DBP in seawater range from 
    0.203 µg/litre in the Kiel Bight (Baltic Sea) (Ehrhardt &
    Derenbach, 1980) and 0.230 µg/litre (Ray et al., 1983a) in Nueces
    Estuary, Texas, up to 4.8 µg/litre in United Kingdom estuaries in
    industrial areas (North and Irish Seas) (Law et al., 1991) and
    24.1 µg/litre in the Baltic and North Seas off the coast of Germany
    (von Westernhagen et al., 1987).

    5.1.2.4  Precipitation

         Atlas & Giam (1981) reported concentrations of DBP in
    rainwater ranging from 0.0026 to 0.0725 µg/litre at the Enewetak
    Atoll in the North Pacific Ocean.  Eisenreich et al. (1981)
    reported that concentrations of DBP in rainwater in the Great
    Lakes area ranged from 0.004 to 0.01 µg/litre; however, no
    sampling or analytical details were given.  In Japan in 1974
    levels of DBP in rainwater ranged from 0.13 to 52 µg/litre
    (detected in 68 out of 111 samples; detection limits ranged from
    0.1 to 4 µg/litre) (personal communication by the Environment
    Agency, Japan, to the IPCS 1995).

         In 1992 DBP concentrations were measured in rainwater
    samples from 3 sites in industrial areas of Germany (LWA 1993). 
    Mean values of 0.8 to 1.4 µg/litre and maximum values of 1.1 to
    4.5 µg/litre were determined.  In woodland areas of Germany that
    are not directly influenced by industry or agriculture, DBP
    concentrations in rainwater were measured at four sites in 1988
    and 1989 (Schleyer et al., 1991).  Outside the forest, mean
    concentrations of 0.21 to 0.35 µg/litre were found.  The
    precipipitation sampled below the trees contained nearly the same
    amount of DBP; at one site the concentration was slightly higher
    with  0.52 µg/litre.  A minimum concentration of 0.06 µg/litre
    and a maximum concentration of 1 µg/litre were found.

    5.1.2.5  Effluent and wastewater

         Concentrations of DBP in effluent ranged from not detectable
    to 61 µg/litre for five Canadian organic chemical plants (number
    of samples unspecified), from not detectable to 94 µg/litre for
    industrial and municipal plants in Cornwall, Ontario (number of
    samples unspecified) and from 1.0 to 100 µg/litre for petro-
    chemical refineries along the St. Clair River (n= 28) (CCREM,
    1987).  The detection limit for this study was 1.0 µg/litre. 
    Concentrations of DBP in fifteen 24-h composite samples of
    process waters collected in 1981 from Canadian refineries
    (unspecified locations) ranged from traces (detection limit,
    2 µg/litre) to 56 µg/litre (PACE, 1985).  However, DBP was not
    detected in 19 samples of effluent discharge of non-chlorinated
    primary-treated municipal wastewater collected in Vancouver in
    1983 (Rogers et al., 1986).

         The concentration in sewage treatment plant effluent from
    Manchester, United Kingdom, sampled during 1984, was 6.0 œg
    DBP/litre (Fatoki & Vernon, 1990).

    5.1.3  Sewage sludge

         DBP has been detected in sludge from municipal wastewater
    plants in Canada (Webber &  Lesage, 1989).  Concentrations ranged
    from 0.2 to 161 mg/kg dry weight in Winnipeg in 1981 and 1982. 
    In Hamilton, the concentrations ranged from 14 mg/kg dry weight
    in 1983 to 57 mg/kg dry weight in 1981.  The authors noted that
    recovery of phthalate esters was erratic, possibly due to
    laboratory contamination or lack of sample homogeneity.

         DBP concentrations were investigated in anaerobic digester
    sludge from nine German municipal wastewater treatment plants
    (Zurmühl, 1990).  In eight plants concentrations were in the
    range of 2.3 to 26 mg/kg dry weight (detection limit =
    1.9 mg/kg). A level of 236 mg/kg dry weight was found as the
    maximum value.  Sewage sludge from another municipal wastewater
    plant contained 0.87 mg DBP/kg dry weight (Kördel & Müller 1992).

    5.1.4  Soil

         DBP levels of < 0.1 to 1.4 µg/g were detected in 13 out of
    30 samples (detection limit, 0.1 µg/g) of soils in urban areas of
    Port Credit and Oakville/Burlington, Ontario (Golder Associates,
    1987).  Concentrations in the background samples on- and off-site
    were similar (Golder Associates, SENES Consultants Limited and
    CanTox, 1987).

         Kördel & Müller (1992, 1993) investigated the DBP
    concentrations in soil in the vicinity of phthalate-emitting
    plants and compared them to a remote area.  There was a great

    deal of variability in the concentrations at the different
    sampling sites, resulting in the fact that no influence of the
    phthalate-emitting plants on soil DBP levels could be derived. 
    The concentrations for the remote site were in the range of <
    0.005 mg/kg to 0.185 mg/kg dry weight.  In the vicinity of the
    industrial sites the values were < 0.005 to 0.560 mg/kg dry
    weight.

    5.1.5  Sediment

         The levels of DBP in sediment are summarized in Table 5.

         Samples of sediment collected from the Detroit River in 1982
    contained concentrations of DBP ranging from < 0.1 to 0.65 mg/kg
    dry weight (Fallon & Horvath, 1985).  Concentrations of DBP in
    sediment samples taken in 1982 from the Fraser Estuary, British
    Columbia, ranged from 0.07 to 0.45 mg/kg dry weight (Rogers &
    Hall, 1987).  The concentration of DBP decreased from 0.204 mg/kg
    dry weight in sediment 0.5 km from a large sewage outfall in the
    estuary to 0.060 mg/kg in sediment 1.0 km from the outfall
    (Rogers & Hall, 1987). Concentrations of DBP up to 0.3 mg/kg were
    reported in samples of sediment collected from Lake Superior and
    Lake Huron in the 1970s (CCREM, 1987).  Concentrations of DBP in
    sediment from the Neckar River in Germany ranged from 0.09 to
    0.3 mg/kg (Malisch et al., 1981).  Higher concentrations (0.028
    to 0.9 mg/kg) were reported in sediment in Maryland, USA
    (Peterson & Freeman, 1984).  Marine sediment from the Crouch
    Estuary United Kingdom contained  0.0039 to 0.0145 mg/kg
    (Waldock, 1983).  Reported concentrations of DBP from marine
    sediments in the USA ranged from 0.0042 mg/kg dry weight in
    Nueces Estuary, Texas (Ray et al., 1983a) to 0.355 mg/kg dry
    weight at Los Angeles (Swartz et al., 1985).  In Japan, levels in
    1974 and 1982 ranged from 0.001 to 2.3 mg/kg (detected in 41 -
    86% of total of 415 samples; detection limits, 0.0007 to
    0.28 mg/kg).

         DBP concentrations in Rhine sediments were measured in 1991. 
    In seven samples concentrations ranged from 0.14 to 2.2 mg/kg dry
    weight.  In 9 out of 10 samples of sediments of the River Weser,
    DBP was detected at concentrations of 0.03 to 0.34 mg/kg dry
    weight with one maximum value of 9.1 mg/kg.  The detection limit
    was 0.02 mg/kg (LWA, 1993).  In Sweden sediment samples from
    different types of enviornment were taken in 1994 (Parkman &
    Remberger, 1995). DBP concentrations in samples from remote sites
    were in the range from 1 to 8 µg/kg dry weight, with one outlier
    of 56 µg/kg (average of three samples per site).  Concentrations
    in industrialized areas were 0 to 182 µg/kg dry weight (detection
    limit = 1.9 µg/kg).

    5.1.6  Aquatic organisms

         In early studies, the concentrations of DBP in aquatic biota
    from the Great Lakes and other areas in Canada were less than

    10 mg/kg (Williams, 1973; Glass et al., 1977; Swain, 1978;
    Burns et al., 1981).  The highest concentrations were reported
    for skinless fillets from long-nose  suckers,  Catostomus
     catostomus, (8.1 µg DBP/g) and rainbow trout,  Oncorhynchus
     mykiss, (5.4 µg/g) from Lake Superior (Glass et al., 1977). 
    In fish from various US Great Lakes harbours and tributary mouths
    in the USA, the concentrations of DBP in the majority of the
    samples ranged from < 0.02 to 0.16 µg/g wet weight; however,
    there were some higher values ranging up to 35 µg/g in more
    polluted areas (DeVault, 1985).  Ray et al. (1983b) reported
    concentrations of DBP in the marine polychaete worm  Neanthes
    virens from Portland, Maine, USA, ranging from 0.070 to
    0.180 mg/kg.

    5.1.7  Terrestrial organisms

         Data on phthalate levels in wild birds and mammals are very
    sparse.  In an early study, Zitko (1972) detected DBP in egg
    yolks of the double-crested cormorant,  Phalacrocorax auritus,
    (14.1 µg/g lipid) and herring gull,  Larus argentatus, (10.9,
    17.1 and 19.1 µg/g lipid).

    5.2  General population exposure

    5.2.1  Ambient air

         Data on concentrations of DBP in ambient air are extremely
    limited.  The most extensive information available is the range
    of concentrations of 4.5 (mean of 15 samples; gas phase) to
    6.2 ng/m3 (mean of 19 samples; particulate phase) in air sampled
    along the Niagara River in 1983 (Hoff & Chan, 1987).  These
    values are similar to those determined more recently in a small
    number of ambient air samples from Barcelona, Spain (Aceves &
    Grimalt, 1993).  Based upon a daily inhalation volume for adults
    of 22 m3, a mean body weight for males and females of 64 kg, the
    assumption that 4 of 24 h are spent outdoors (IPCS, 1993) and the
    above range of concentrations in ambient air, the mean intake of
    DBP via ambient air for the general population is estimated to
    range from 0.26 to 0.36 ng/kg body weight per day.

    5.2.2  Indoor air

         The maximum concentration of DBP in indoor air in nine homes
    in Montreal, Canada, sampled for three consecutive periods of 20
    days each, was 2.85 µg/m3 (nominal quantification limit,
    0.50 µg/m3) (Otson & Benoit, 1985).  No other information on
    measured concentrations (e.g., mean concentrations) was
    presented.  In a survey of 125 homes in California in 1990, the
    median daytime concentration of DBP in indoor air was 420 ng/m3
    (California Environmental Protection Agency, 1992).

         Based upon a daily inhalation volume for adults of 22 m3, a
    mean body weight for males and females of 64 kg, the assumption

    that 20 of 24 h are spent indoors (IPCS, 1993) and the median
    concentration of DBP reported in a survey of a large number of
    homes in California (420 ng/m3), the daily intake of DBP in
    indoor air for the general population is estimated to be
    120 ng/kg body weight per day.

    5.2.3  Drinking-water

         Data on concentrations of DBP in drinking-water are limited.
    In an early survey (1974), DBP was detected (detection limit
    unspecified) in six out of ten city water supplies in the USA. 
    The concentrations of DBP ranged from 0.01 to 0.1 µg/litre for
    five cities and was 5.0 µg/litre for one city (Keith et al.,
    1976).  Concentrations in two samples of tap water from the
    Shizuoka Prefecture in Japan taken in 1974 were 1.0 and
    0.8 µg/litre (Shibuya, 1979).  In samples of tap and well water
    in Japan, levels were 1.9 and 2.5 µg/litre, respectively (Ishida
    et al., 1980).  In a survey of an unspecified number of samples
    of the municipal drinking-water supplies of seven cities in the
    Niagara region and in the vicinity of Lake Ontario conducted in
    1984 (MOE, 1984), DBP was not detected (detection limit,
    1.0 µg/litre).

         In a small number of samples of drinking-water in Toronto,
    Canada, the mean concentration was 14 ng/litre; concentrations in
    seven brands of bottled spring water ranged from 21 to
    55 ng/litre (City of Toronto, 1990).

         Based upon a daily water consumption for adults of 1.4
    litres, a mean body weight for males and females of 64 kg (IPCS,
    1993) and a mean concentration of < 1.0 µg/litre, the estimated
    mean intake of DBP from drinking-water for the general population
    is <0.02 µg/kg body weight per day.

    5.2.4  Food

         In addition to entry through environmental contamination,
    DBP may be present in foodstuffs as a result of migration from
    packaging.  This has been investigated in a number of studies
    conducted in the late 1980s.  In many countries, on the basis of
    the results of these studies, precautions were introduced to
    reduce leaching of plasticizers from packaging.  As a result,
    levels of DBP in foodstuffs have declined over time.  In this
    section, studies designed to investigate the presence of DBP in
    foodstuffs due to leaching from packaging are presented, followed
    by data from more broadly based  market-basket surveys.

         Concentrations of DBP ranged from 0.13 to 1.62 mg/kg in
    three brands of aluminum foil in Japan (Ishida et al., 1980).

         In the first of several studies conducted in the United
    Kingdom to investigate the impact of packaging on the DBP content
    of foodstuffs, foods were purchased at retail stores and stored
    in their packaging until their "sell by" or "best before" date
    (British Ministry of Agriculture, Fisheries and Food, 1987). 
    Mean concentrations of DBP were 8 to 32 mg/kg in chocolate
    confectionery, 13 mg/kg in sugar confectionery, 11 mg/kg in
    cakes, 3.9 to 11 mg/kg in baked savouries, 6 to 10 mg/kg in meat
    pies and 2 mg/kg in sandwiches.

         In a survey of plastic-packaged Italian foodstuffs, DBP was
    detected in cheese (0.84 œg/g), salted meat (1.09 mg/kg),
    vegetable soups (2.06 mg/kg), potato chips (2.80 mg/kg) and
    pasteurized milk (0.07 mg/kg) (Cocchieri, 1986).

         Levels of DBP ranged from 0.5 to 30.8 mg/kg in nougat and
    chocolate, respectively, in a wide range of foodstuffs in the
    United Kingdom, which were wrapped in a range of different
    packaging including nitrocellulose-coated regenerated cellulose
    film (RCF). Levels of plasticizers were 0.5 to 1.5%, on a total
    film-weight basis (Castle et al.,1988).  In a later study, Castle
    et al. (1989) reported that DBP in the ink on the outer surface
    of film can transfer onto the inner food contact surface.  The
    level of DBP in a chocolate-covered confectionery product
    increased from 0.2 to 6.7 mg/kg over a storage period of 180
    days.  DBP levels in 47 samples of confectionery, snack products
    and biscuits purchased in the United Kingdom, wrapped in printed
    polypropylene film, ranged from 0.02 to  14.1 mg/kg.

           In a more recent reported retail survey in the United
    Kingdom (MAFF, 1990), ranges in up to 30 samples each of plastic
    wrapped foods were 0.09 to 0.13 mg/kg in biscuits, 0.02 to
    14.1 mg/kg in potato snacks, 0.15 to 5.6 mg/kg in chocolate-
    covered bars and 2.6 to 9.2 mg/kg in candy-coated chocolate
    sweets.  In the same report, results of sequential analysis of a
    few foods were also reported.  Concentrations in potato snacks,
    candy-coated individual sweets and chocolate bars increased
    approximately 2- to 3-fold over a 6-month period.

         Page & Lacroix (1992) reported that retail samples of
    packaged butter and margarine sold in Canada contained up to
    10.6 mg DBP/kg.

         Nerin et al. (1993) analysed plastic-wrapped food products
    for DBP from both Spain and the United Kingdom and reported (for
    an average of three determinations) up to 0.81 mg/kg in chocolate
    bars and 0.60 mg/kg in biscuits.

         In an early Canadian study (Williams, 1973), DBP was
    determined in 21 samples of fish.  DBP was detected in one sample
    of canned tuna at a concentration of 78 µg/kg while the levels in
    one  sample of canned salmon was 37 µg/kg.  Concentrations of DBP
    in the muscle of fish (n = 10 samples from five species) from the
    lower Fraser River in British Columbia ranged from 0.07 to
    0.15 mg/kg wet weight (Swain & Walton, 1989).  The authors
    considered 0.07 mg/kg as the background level, owing to
    contamination; the detection limit was not reported.  Elevated
    concentrations of DBP have occasionally been reported in fish in
    polluted areas (see section 5.1).

         Based upon residue analysis of commercial eggs collected
    throughout Japan, 0.098 mg DBP/kg  (trace - 0.15 mg/kg was
    present in egg whites (Ishida et al., 1981).  No phthalate
    residues were found in the egg yolks. In an early study of 2 to
    14 samples each of various foodstuffs in Japan, DBP was detected
    in meat (100 µg/kg), fish (180 µg/kg), eggs (80 µg/kg), but not
    in milk (detection limit, 50 µg/kg) (Howard, 1989).  In another
    study (Tomita et al., 1977), DBP was determined by gas-liquid
    chromatography (detection limit, 0.01 mg/kg) in 22 kinds of
    Japanese foods (17 samples of fatty foods and 38 samples of non-
    fatty foods mostly in plastic containers).  DBP was detected in
    tempura (frying) powder (0.39 to 17.70 mg/kg), instant cream soup
    (1.73 to 60.37 mg/kg), fried  potato cake (not detected to
    1.11 mg/kg), orange juice (0.35 mg/kg) and pickles (0.11 mg/kg).

         Ito et al. (1993) reported that 2 out of 15 samples of
    imported vodka in Japan contained up to 0.2 mg DBP/litre. In the
    USA, DBP was detected in 18 out of 50 samples of vodka (maximum
    concentration: 204 µg/litre; limit of detection: 20 µg/litre)
    (Leibowitz et al., 1995). DBP was detected in 1 out of 60 samples
    of Russian vodka (0.7 mg/litre) and in 1 out of 7 samples of
    European vodka (1.1 mg/litre) (Saito et al., 1993).

         In a Canadian market-basket survey of 98 different food
    types sampled in Halifax in 1986 (Page & Lacroix, 1995), DBP was
    detected in butter (1.5 mg/kg), freshwater fish (0.5 mg/g),
    cereal  products (ranged from not detected to 0.62 mg/kg), baked
    potatoes (0.63 mg/kg), coleslaw (0.11 mg/kg), bananas,
    blueberries and pineapples (0.12, 0.09 and 0.05 mg/kg,
    respectively), margarine (0.64 mg/kg), white sugar (0.2 mg/kg)
    and gelatin dessert (0.09 mg/kg).  The detection limits varied
    (ranging from 0.01 to 0.5 mg/kg) according to the reagent blank
    values (interferences arising from coextracted food components)
    and the fat content of the food.

         Exposure of the general population to DBP in food has been
    estimated on the basis of data from the only study identified in
    which there was a sufficiently wide variety of foodstuffs to
    serve as a basis, i.e., those from a market-basket survey in
    Canadaa.  Based upon the average daily consumption of various
    foodstuffs by adultsb, a mean body weight for males and females
    of 64 kg (IPCS, 1993) and concentrations of DBP reported in the
    Canadian market basket survey, the estimated daily intake from
    food is 7 µg/kg body weight per day.  It should be noted,
    however, that intake of DBP in the diet can vary considerably,
    depending upon the nature and amount of packaged food that is
    consumed and the nature of use of food wrapping in food
    preparation.  In the United Kingdom, the Ministry of Agriculture,
    Fisheries and Food has estimated that the maximum likely human
    intake of DBP from food sources is approximately 2 mg per person
    per day (approximately 31 œg/kg body weight per day, assuming a
    mean body weight of 64 kg).

    5.2.5  Consumer products

         In 1981, DBP was reported as an ingredient in a total of 590
    cosmetic formulations in the USA, at concentrations ranging from
    less than 0.1% to between 10 and 25% (Brandt, 1985).  There is
    potential for exposure to DBP in cosmetics, but available data
    are inadequate to quantify intake from this source.

         The "new car smell" in automobiles has been attributed to
    DBP and other phthalic acid esters (Shea, 1971).  Levels of total
    phthalic acid esters in the µg/m3 range have been identified in
    samples of air taken from new cars in an early study (Graham,
    1973).

    5.2.6  Medical devices

         Plastic tubing used in hospitals for oral/nasal feeding of
    patients, has been reported to contain 54 mg DBP/g (Khaliq et

              

    a    Data from the Canadian market-basket survey used in
         calculating the estimated average daily intake include
         concentrations of DBP in the following foodstuffs: butter,
         1.5 mg/kg; freshwater fish, 0.5 mg/kg; cereal products,
         0.62 mg/kg, baked potatoes, 0.63 mg/kg; bananas, 0.12 mg/kg;
         white sugar, 0.2 mg/kg.

    b    Dietary intakes consist of: cereals, 323 g/day; starchy
         roots, 225 g/day; sugar (excludes syrups and honey),
         72 g/day; pulses and nuts, 33 g/day; vegetables and fruits,
         325 g/day; meat, 125 g/day, eggs, 19 g/day; fish, 23 g/day;
         milk products (excludes butter), 360 g/day; fats and oils
         (includes butter), 31 g/day (IPCS, 1993).

    al., 1992).  DBP leached from tubing into distilled water and
    solutions of ethanol, acetic acid and sodium bicarbonate, in
    concentrations which increased with temperature and duration of
    contact.

    5.2.7  Levels in human tissue

         In an early study, concentrations of DBP in 25 samples of
    human adipose tissue collected from Vancouver (n = 2), Toronto (n
    = 22) and Montreal (n = 1) at autopsies of accident victims,
    ranged from 0.01 to 0.3 mg/kg (detection limit not reported) (Mes
    et al., 1974).

         Levels of DBP in the blood collected from 13 individuals
    (mean, 0.10 mg/litre) following ingestion of food that had been
    in contact with unspecified flexible plastics packaging materials
    containing DBP were higher than those collected from nine
    individuals before meals (mean levels in blood, 0.02 mg/litre)
    (Tomita et al., 1977).

    5.3  Occupational exposure

         Identified data on levels of DBP in the occupational
    environment are limited.  Based on a survey conducted by the
    National Institute of Occupational Safety and Health (NIOSH) in
    1981-1983, it was estimated that there were 229 000 workers in
    the USA with potential exposure to DBP (Howard, 1989). The most
    recent provisional data from the National Occupational Exposure
    Survey indicates that over 500 000 workers, including 200 000
    women are potentially exposed to DBP (NIOSH, 1994).

         In 1986, NIOSH conducted a health hazard evaluation of a
    silkscreening area in a Department of Highways sign shop (NIOSH,
    1987).  Concentrations of DBP were below the limit of detection
    (less than 0.01 mg per sample), i.e., less than 0.02 mg/m3.

         Only trace quantities of DBP were detected in a 1975 survey
    of a Goodyear Tire and Rubber Company plant in areas involved in
    the production of rubber sleeve stock (NIOSH, 1976).

         In 1981, an environmental survey was conducted at a US army
    ammunition plant, in an area where DBP-containing propellant was
    processed (NIOSH, 1982).  Four samples (1 breathing zone, 3 area)
    were collected.  One area sample contained DBP in an amount
    corresponding to a concentration of 0.08 mg DBP/m3.  The other
    three samples contained less than the detection limit
    (0.01 mg/sample).

         An industrial hygiene survey was conducted in a plastic pipe
    fabricating plant in the USA in 1988.  Six personal breathing
    zone air samples collected for DBP were below the level of
    detection,  corresponding to < 0.01 mg/m3 (NIOSH, 1989).

         Fischer et al. (1993) reported that concentrations of DBP
    ranged from 1.3 to 8.2 mg/m3 in a plant in the Czech Republic
    that produced PVC products.

         Thus, based on determinations at a limited number of
    worksites in the USA, concentrations have generally been less
    than the limit of detection (i.e., 0.01 to 0.02 mg/m3), although
    levels of up to 8 mg/m3 were reported in a PVC plant in the
    Czech Republic.

    6.  KINETICS AND METABOLISM IN LABORATORY ANIMALS AND HUMANS

         Data on kinetics and metabolism in mammals are presented in
    this chapter.  Information on metabolism in invertebrates is
    presented in Chapter 4.

    6.1  Absorption, distribution and excretion

    6.1.1  Dermal

         A study was conducted by Elsisi et al. (1989) in which
    157 µmol/kg (43.7 mg/kg) of 14C-DBP (uniformly labelled on the
    ring) was applied to the back of male F-344 rats and the area of
    application was covered with a perforated cap for a 7-day
    period).  Approximately 10 to 12% of the administered dose was
    excreted in the urine each day for several days (total of 60%
    after 1 week).  Only small amounts of radioactivity were detected
    in tissues in the exposed rats.  About 33% of the dose remained
    at the site of application; all other tissues combined contained
    less than 0.5% of the applied dose.

         Based on results observed  in vitro, Scott et al. (1987)
    reported that DBP was slowly absorbed through both rat and human
    skin, with rat skin being more permeable.

    6.1.2  Ingestion

    6.1.2.1  In vivo studies

         Levels of DBP in the blood collected from 13 individuals
    (mean, 0.10 mg/litre) 2 h following ingestion of food, which had
    been in contact with unspecified flexible plastic packaging
    materials containing DBP, were higher than those collected from
    nine individuals before meals (mean level in blood,
    0.02 mg/litre) (Tomita et al., 1977).

         Studies in experimental animals indicate that DBP or its
    metabolites are rapidly absorbed from the gastrointestinal tract. 
    In a study conducted by Williams & Blanchfield (1975), following
    administration of a single oral dose of about 0.1 g/kg body
    weight 7-14C-DBP to male Wistar rats, 96% of the radioactivity
    was excreted in the urine at 48 h; less than 0.1% was exhaled as
    14CO2.  In addition, blood and tissue levels and urine output
    were determined at 4, 8, 24 and 48 h following administration of
    single oral doses of 7-14C-DBP (0.27 or 2.31 g/kg body weight). 
    The  radioactivity was distributed more or less evenly throughout
    the tissues except that the level in the brain was about one
    third to one tenth that in the other tissues.  Excretion in the
    urine was rapid, with 46% of the low dose and 20% of the high
    dose being present in the urine at 8 h, 85 and 61%, respectively,
    at 24 h, and 92 and 83%, respectively, at 48 h.  Based on
    analysis of the urine, 80 to 90% of the dose was metabolized and

    excreted in the urine in 48 h as phthalic acid (2%), mono-
     n-butyl phthalate (88%), mono 3-hydroxy butyl phthalate (8%)
    and mono-4-hydroxy butyl phthalate (2%).  These authors also
    reported that there was no evidence of accumulation in any
    tissues in rats fed 0.1% DBP in the diet for 4, 8 or 12 weeks.

         Twenty four hours following gavage (in 3% DMSO solution)
    administration of a single dose of 60 mg/kg body weight 14C-DBP
    to small groups (n=3) of male Wistar rats, radioactivity was
    detected in the liver, kidney, blood, muscle, adipose tissue,
    stomach and intestine (the latter probably associated with
    biliary excretion).  There was no significant retention of DBP
    within tissues; more than 90% of the administered radioactivity
    was recovered in the urine within 48 h (Tanaka et al., 1978).

         In DSN hamsters, 79% of a single oral dose of 2 g/kg body
    weight (10 µCi of 14C-DBP/kg body weight) administered by gavage
    was excreted in the urine within 24 h, mainly as mono- n-butyl
    phthalate (Foster et al., 1982).

    6.1.2.2  In vitro studies

         Mono- n-butyl phthalate (MBP) was absorbed in significantly
    greater quantity than DBP in an  in vitro study in an everted
    gut-sac preparation from the small intestine of male Sprague
    Dawley rats (White et al., 1980).  DBP was actively hydrolysed by
    esterases within the mucosal epithelium during absorption; 95.5%
    of DBP was hydrolysed to MBP.  When the esterase activity of the
    mucosa was reduced by intragastric exposure of the rats to S,S,S-
    tributylphosphorotrithioate (8 mg/kg body weight), the absorption
    of DBP, but not of MBP, was significantly reduced (from 0.62 to
    0.15 µmol/mg per h).

    6.1.3  Inhalation

         Following inhalation by rats of 50 mg/m3  for various
    periods up to 6 months (Kawano, 1980b), DBP was detected by GC/MS
    at relatively low concentrations in the brain (0.53 µg/g), lung
    (0.17 µg/g) and liver (0.25 µg/g) of small groups of male Wistar
    rats.  Levels in the testes were lower (mean 0.13 œg/g).
    Following exposure to 0.5 mg/m3 (0.044 ppm), DBP was
    consistently detected only in the brain of exposed rats.

    6.2  Metabolic transformation

    6.2.1  In vivo studies

         Available data indicate that in rats DBP is metabolized by
    nonspecific esterases, mainly by hydrolysis, to yield MBP, with
    subsequent oxidation of the alkyl side chain of MBP. 
    Interestingly, MBP is stable and resistant to hydrolysis of the

    second ester group (Cater et al., 1977; Rowland et al., 1977). 
    Following oral administration of DBP to rats, metabolic products
    identified in the urine were mainly MBP, various oxidation
    products of MBP (2-3%), and a small amount of the free phthalic
    acid (Albro & Moore, 1974; Williams & Blanchfield, 1975; Foster
    et al., 1982).  The MBP and other metabolites are excreted in the
    urine mainly as glucuronide conjugates; species differences in
    the excretion of conjugated and unconjugated metabolites of DBP
    in the urine of Wistar rats and DSN hamsters have been observed.
    In hamsters, 53% was excreted as the conjugate and 3.5% as free
    monoester.  In rats, 38% was excreted as conjugate and 14% as
    free monoester, following administration of an oral dose of 2
    g/kg body weight (10 µCi of 14C-DBP/kg body weight per day) by
    gavage.  No free DBP was detected in the urine in either species
    (Foster et al., 1982).

    6.2.2  In vitro studies

         In  in vitro studies, DBP was hydrolysed to MBP by cell
    preparations from the small intestine (rat, baboon, man), the
    liver (rat, baboon) and kidneys (rats) (Lake et al., 1977; Tanaka
    et al., 1978; Kaneshima et al., 1978).

         Rowland et al. (1977) incubated the contents of the male
    Wistar rat stomach, small intestine and caecum with 14C-labelled
    DBP for 16 h.  About 0.5, 80 and 23% of the DBP was hydrolysed to
    MBP by the contents of the stomach, small intestine and caecum,
    respectively.  The metabolism of DBP by the small intestinal
    contents was very rapid, 38% of a dose of 1 mg DBP/ml and 70% of
    a dose of 200 œg/ml being metabolized in 30 min.  Thus, it would
    appear that DBP is relatively quickly converted to MBP in the
    intestines, this being the principal metabolite.  Activity in the
    female rat small intestine was only slightly less than that for
    the male.  Suspensions prepared from human faeces also had modest
    DBP hydrolytic activity (6% in 16 h) (Rowland et al., 1977). 
    Because activity did not decrease when antibiotics were present
    during the incubation, the author concluded that the enzymatic
    hydrolytic activity was of mammalian origin (possibly pancreatic
    and mucosal lipases).

         Using 14C-DBP as substrate, the rate of esterase activity
    was comparable in small intestinal tissue of rats and hamsters,
    whereas the liver of hamsters had approximately double the
    activity of rats.  In contrast, the ß-glucuronidase activity of
    testicular homogenates in the rat was much higher than that in
    the hamster ( p-nitrophenyl glucuronide and phenolphthalein
    glucuronide were used as substrates) (Foster et al., 1982).

         In  in vitro assays of rat liver, kidney, pancreas, small
    intestine and blood, structural analogues of DBP (di- n-butyl
    isophthalate and di- n-butyl terephthalate) were hydrolysed to

    their corresponding acids, whereas phthalic acid was not formed
    from DBP (Takahashi & Tanaka, 1989).  The authors  concluded that
    nonionic esters are hydrolysed at a much higher rate than charged
    analogues and that esterase activities are strikingly different
    for different substrates.

    7.  EFFECTS ON LABORATORY MAMMALS AND IN VITRO TEST SYSTEMS

    7.1  Single exposure

         The acute toxicity of DBP in mice and rats is low.  Reported
    LD50 values following oral administration to rats range from
    approximately 8 g/kg body weight to at least 20 g/kg body weight
    (Smith, 1953; Lehman, 1955; White et al., 1983; Brandt, 1985); in
    mice,  values are approximately 5 to 16 g/kg body weight
    (Woodward, 1988; Brandt, 1985; Yamada, 1974).  Reported  LD50
    values following intraperitoneal administration range from 4 to
    7 g/kg body weight in rats and approximately 3 to 6 g/kg body
    weight in mice (Woodward, 1988).  The dermal LD50 in rabbits is
    > 4000 mg/kg body weight (Lehman, 1955).  Signs of toxicity
    include general depression of activity, laboured breathing and
    lack of coordination.  Reports of acute toxicity of DBP following
    inhalation have not been identified.

         Following intraperitoneal injection, MBP (the principal
    metabolite of DBP) appeared to be somewhat more acutely toxic
    than DBP; the LD50 was 1.0 g/kg in the mouse (Chambon et al.
    1971).

    7.2  Short-term exposure

         The short-term toxicity of DBP has been investigated in
    rodents following oral administration.  The available data are
    summarized in Table 6.

         In most of these studies, animals were exposed to only one
    dose level.  Effects in rats after oral administration for 5 to
    21 days include those on liver enzymes (Aitio & Parkki, 1978;
    Bell et al., 1978; Kawashima et al., 1983; BIBRA, 1986; Barber et
    al., 1987) and hepatomegaly at doses of >420 mg/kg body weight
    per day (Yamada, 1974; Bell et al., 1978; Oishi & Hiraga, 1980a;
    BIBRA, 1986; Barber et al., 1987), a reduction in the rate of
    weight gain at doses of >5 ml/kg body weight per day
    (5235 mg/kg body weight per day) (Yamada, 1974) and splenomegaly
    after intragastric intubation of 1.0 ml/kg body weight per day
    (1047 mg/kg body weight per day) (Yamada,  1974).  Peroxisome
    proliferation, based on increased oxidation of cyanide-
    insensitive CoA oxidation, in the liver of male F-344 rats was
    observed after administration of 2100 mg/kg body weight per day
    in the diet for 21 days (Barber et al., 1987) and also in male
    Wistar rats after exposure for 34 to 36 days to 2500 mg/kg body
    weight per day in the diet (Murakami et al., 1986a).
    Proliferation at lower levels has also been reported in an
    investigation summarized in an abstract by Lake et al. (1991).  A
    slight but insignificant increase in kidney weight was reported
    in JCL:Wistar rats exposed to 2060 mg/kg body weight per day for
    7 days by Oishi & Hiraga (1980a).

        Table 6.  Short-term repeated dose toxicity of DBP

                                                                                                                                           

    Species                 Protocol                               Results                              Effect Levels       Reference
                                                                                                                                           

    Rat (Wistar,       1047 or 5235 mg/kg     The rate of b.w. gain was slightly reduced at the high    LOAEL = 1047      Yamada (1974)
    groups of 5        b.w. per day by        dose.  One rat administered the high dose died during     mg/kg b.w.
    females)           stomach tube daily     the study.  Hepatomegaly and marked splenomegaly noted    per day
                       for 3 weeks.           at necropsy in both exposed groups;  relative kidney
                       Controls were          weight of high-dose group 76% greater than that in
                       administered           controls.
                       10 ml/kg distilled
                       water in the same
                       manner.

    Rat (Wistar,       2% DBP in the diet     Marked increases in stearoyl-CoA desaturation,            One dose group    Kawashima
    groups of          (equivalent to 1000    palmitoyl-CoA oxidation and catalase activity;            only (effects     et al. (1983)
    3 males)           mg/kg b.w. per day)    increases in microsomal octadecanoic acid in liver,       observed at
                       for 7 days             hepatic homogenates and serum. The increases in the       1000 mg/kg b.w.
                                              stearoyl-CoA desaturation appeared to be due to the       per day)
                                              increased activity (4 fold) of the terminal
                                              desaturase and not to increases in the activities
                                              of NADH cytochrome-C-reductase or in cytochrome b5
                                              content.

    Rat (JCL:Wistar,   2% DBP in the diet     Mean b.w.s of exposed rats were slightly but not          One dose group    Oishi & Hiraga
    groups of 10       equivalent to 2060     significantly lower than that of the controls.            only (effects     (1980a)
    males)             mg/kg b.w. per day     Significant decrease in absolute and relative             observed at
                       for 7 days             testicular weights, but the absolute and relative         2060 mg/kg b.w.
                                              liver weights were significantly increased.               per day)
                                              Slight but insignificant increase in kidney weight
                                              in exposed rats.
                                                                                                                                           

    Table 6.  Continued

                                                                                                                                           

    Species                 Protocol                               Results                              Effect Levels       Reference
                                                                                                                                           

    Rat (Fischer-344,  dietary                Males at mid and high dose and females at high dose       LOEL = 624        BIBRA (1986),
    5 animals per      administration for     gained less weight than controls.  Absolute and           mg/kg b.w.        Barber et al.
    sex per dose)      21 days at levels      relative liver weight increased in all exposed            per day           (1987)
                       of 0, 0.6%, 1.2%       groups.  Lower testis weight in high-dose males;
                       or 2.5% DBP;           severe atrophy observed upon histopathological
                                              examination.  Serum triglyceride and cholesterol
                       a positive control     levels decreased in all exposed males and cholesterol
                       group was              level reduced in all exposed females, in a
                       administered 1.2%      non-dose-related manner. Slight reduction in
                       di(2-ethylhexyl)       hepatocyte cytoplasmic basophilia in all rats at
                       phthalate;             highest doses and in males at 1.2%.
                                              Cyanide-insensitive palmitoyl CoA oxidation
                       dose levels            increased in both sexes at the highest dose and at
                       (calculated by         the 1.2% dose in males.
                       investigators and      Lauric acid 11 and 12 hydroxylase activities were
                       presented in BIBRA     significantly increased in all exposed males and
                       (1986));               in females in the high-dose group.

                       males:  0, 624,
                       1234, 2156 mg/kg
                       b.w. per day

                       females:  0, 632,
                       1261, 2107 mg/kg
                       b.w. per day
                                                                                                                                           

    Table 6.  Continued

                                                                                                                                           

    Species                 Protocol                               Results                              Effect Levels       Reference
                                                                                                                                           

    Rat (F-344,        0.05, 0.1, 0.5, 1.0    A dose-related liver enlargement and induction of         NOAEL = 104       Lake et al.
    male, groups       or 2.5% DBP in the     palmitoyl-CoA oxidation activity were reported.           mg/kg b.w.        (1991) (abstract)
    of 5 males)        diet for 28 days       Based on the enzyme activity, the no-effect level for     per day
                       (not possible to       induction of hepatic peroxisome proliferation was
                       present doses on a     determined to be 104 mg/kg b.w. per day by the authors.
                       b.w. basis since
                       food consumption was
                       determined but not
                       reported)

    Rat                0.7% DBP in the diet   Hepatomegaly was noted in exposed rats.  Reduction        One dose group    Bell et al.
    (Sprague-Dawley,   (equivalent to 420     in serum cholesterol levels in exposed animals and        only (effects     (1978)
    groups of 9        mg/kg b.w. per day)    inhibition in hepatic sterologenesis reducing the         observed at 420
    males)             for 21 days            uptake of 14C-mevalonate and 14C-acetate by the liver     mg/kg b.w.
                                              minces of the exposed rats.  There was no effect on       per day)
                                              hepatic cholesterol levels.

    Rat (Wistar,       5 mmol/kg b.w. per     Increases in hepatic cytochrome P-450 levels and          One dose group    Aitio & Parkki
    groups of 7        day (1390 mg/kg b.w.   in the activities of epoxide hydratase and                only (effects     (1978)
    males)             per day) in corn oil   glutathione-S-transferase.  No statistically              observed at
                       by gavage for 6 days   significant increase in the catalytic activities          1390 mg/kg b.w.
                                              dependent on cytochrome P-450 (ethoxycoumarin             per day)
                                              de-ethylation and benzo(a)pyrene hydroxylation).

    Mouse (ICR,        2% DBP in the diet     Food consumption was affected (data not presented)        One dose group    Oishi & Hiraga
    groups of 10       (2400 mg/kg b.w.       and b.w. gain was significantly decreased.  The           only (effects     (1980b)
    males)             per day) for 1 week    relative liver weight was significantly increased         observed at
                                              whereas the relative kidney weight was significantly      2400 mg/kg b.w.
                                              reduced.  The zinc concentration in the liver was         per day)
                                              reduced to 88 ± 3.14% of the control value while
                                              that of the kidney remained unchanged.
                                                                                                                                           
             For mice, identified data on short-term toxicity are limited
    to one investigation, in which there was a significant decrease
    in the relative kidney weight when ICR male mice were fed a diet
    containing 2% (equivalent to 2400 mg/kg body weight per day) DBP
    for 1 week (Oishi & Hiraga, 1980b).  Results of histopathological
    examinations were not reported.

         In a short-term study, for which only an abstract was
    published, Lake et al. (1991) compared the relative potentials of
    several phthalates, including DBP and DEHP, to induce peroxisome
    proliferation and testicular atrophy in rats.  DEHP was more
    potent than DBP in inducing palmitoyl-CoA oxidation activity. 
    The NOEL values for induction of peroxisome proliferation
    activity were  considered to be 52 and 104 mg/kg body weight per
    day for DEHP and DBP, respectively.  In  contrast, the values for
    testicular atrophy were 1093 and 515 mg/kg body weight per day
    for DEHP and DBP, respectively.

         In another study reported by Barber et al. (1987), in which
    groups of five male and five female rats were administered 2.5%
    DBP in the diet (1250 mg/kg body weight per day) for 21 days, the
    extent of peroxisome proliferation in the males, on the basis of
    electron micrographs, was equivalent to that produced by 0.6%
    DEHP (300 mg/kg body weight per day).  A scale of peroxisome
    proliferation activity in the male rat was drawn up by the
    investigators based on their own short-term results and work
    published by other investigators.  Relative values, based on
    these studies conducted by various investigators for fenofibrate,
    ciprofibrate, Wy 14,643, DEHP, DBP and aspirin were 304, 66, 44,
    15, 3 and 1, respectively.

    7.3  Long-term exposure

         The effects of long-term exposure to DBP have been 
    investigated in several studies on rodents following oral
    exposure; however, only limited information concerning effects
    following inhalation was identified.  Details of study design and
    results are presented in Table 7.

         In the study by Nikonorow et al. (1973), groups of 20 Wistar
    rats (10 males and 10 females) were administered 120 and
    1200 mg/kg body weight per day for 3 months by gavage in olive
    oil.  At both doses, there was a statistically significant
    increase in the relative liver weight.  No particular 
    alterations in the liver, kidneys and spleen of any rats
    administered DBP were seen during gross or  histological
    examination.  The LOEL was considered to be 120 mg/kg body weight
    per day, based  on the increase in relative liver weights.

         A series of three subchronic dietary studies have been
    published recently (NTP, 1995):

        Table 7.  Long-term toxicity of DBP

                                                                                                                                            

    Species                   Protocol                                      Results                         Effect Levels      Reference
                                                                                                                                            

    Oral

    Rat (Wistar,      120 or 1200 mg/kg b.w. per        Clinical signs of toxicity were not described.      LOEL = 120         Nikonorow et
    groups of 10      day in olive oil by gavage        One rat from the high-dose group died but the       mg/kg b.w.         al. (1973)
    males and 10      daily for 3 months                death was not considered to be treatment-related    per day
    females)                                            (sex not specified).  At necropsy, the only
                                                        effect noted was hepatomegaly at both doses.
                                                        No gross or microscopic changes were noted in
                                                        the spleen or kidneys.

    Rat (strain       2300 mg/kg b.w. per day           There was a reduction in the rate of weight         one dose group     Radeva &
    and sex           administered by gavage            gain from day 1 onwards, but no other               only (reduction    Dinoyeva
    unspecified,      (vehicle unspecified) for         clinical signs of oxicity were noted and no         in weight gain     (1966)
    groups of 8;      50 days                           deaths occurred.  No other end-points were          at 2300 mg/kg
    5 as controls)                                      reported.                                           b.w. per day)


    Rat (strain       diets containing levels           There were no clinical signs of toxicity and        cannot be          Radeva &
    not specified,    equivalent to 0.1, 1 and          no haematological abnormalities.  Urine             determined         Dinoyeva
    groups of 8       10 mg/kg b.w. per day for         analyses for hippuric acid, albumin and                                (1966)
    males)            7 months;  controls               sediment contents were normal.  Marked
                      received the vehicle used         venous congestion in some exposed rats at
                      to produce the feed mix,          necropsy was reported, but the organ and
                      sunflower oil, in the diet        dose group(s) in which it occurred were not
                                                        specified.  No other compound-related lesions
                                                        were noted.
                                                                                                                                             
    

    Table 7.  Continued

                                                                                                                                             
    

    Species                   Protocol                                      Results                         Effect Levels      Reference
                                                                                                                                             
    

    Rat (Wistar,      5% DBP in the diet                Reduction in b.w. gain during the first week,       one dose group     Murakami
    groups of 5       (equivalent to a dose of          followed by a plateau, which was 65% of the         only (effects      et al.
    males)            2500 mg/kg b.w. per day)          control value on the 35th day.  There were          observed at 2500   (1986b)
                      for 35 to 45 days                 significant increases in relative liver and         mg/kg b.w. per
                                                        spleen weights but no significant changes in        day)
                                                        absolute weight.  In addition, there was
                                                        marked atrophy of the testicles.  There was
                                                        also depressed respiration in liver
                                                        mitochondria when succinate or pyruvate was
                                                        used as a substrate.  Hepatic glutamate
                                                        dehydrogenase activity was decreased by 73%
                                                        of the control value but this was not
                                                        significant.  Succinate and pyruvate
                                                        dehydrogenase activities were significantly
                                                        decreased (by 59% and 38% of the control
                                                        values, respectively).
                                                                                                                                             
    

    Table 7.  Continued

                                                                                                                                            

    Species                   Protocol                                      Results                         Effect Levels      Reference
                                                                                                                                            

    Rat (Wistar,      0.5% or 5% in the diet            B.w., expressed as percentage of weight in          LOAEL = 250        Murakami
    groups of 5       (equivalent to 250 or             the control group, decreased gradually in           mg/kg b.w.         et al.
    males)            2500 mg/kg b.w. per day,          both groups.  There were significant                per day            (1986a)
                      respectively) for 34 to           increases in the relative weights of the
                      36 days                           liver, kidney and spleen, and decreases in
                                                        the weight of the testicles in the high-dose
                                                        group.  The succinate and pyruvate
                                                        dehydrogenase activities in liver
                                                        mitochondria were significantly inhibited at
                                                        the high dose, but not the glutamate
                                                        dehydrogenase activity.  The activities of
                                                        AP, GOT and GPT increased in rats that
                                                        received the high dose.  Decreased globulin
                                                        and increased albumin/globulin ratio were
                                                        observed in both dose groups.  In the
                                                        liver, cytotoxic injury including single
                                                        cell necrosis, zonal necrosis and degeneration
                                                        with ballooning were observed in many of the
                                                        rats that received the high dose.  Zonal
                                                        necrosis and liver atrophy were observed in
                                                        the low-dose group.  Ultrastructural
                                                        examination of liver cells revealed that the
                                                        effects of exposure to the high dose were more
                                                        extensive than the low dose in increasing
                                                        peroxisomes, lysosomes and mitochondria.
                                                                                                                                            

    Table 7.  Continued

                                                                                                                                            

    Species                   Protocol                                      Results                         Effect Levels      Reference
                                                                                                                                            

    F-344 rat;        This study has been designated    Final b.w.:  reduction in males at >10 000          LOEL = 356         NTP (1995)
    10 per sex        "NTP Study 2"; administration     mg/kg and in females at >20 000 mg/kg               mg/kg b.w.
    per group         in diet for 13 weeks at 0,        Organ weights:  hepatomegaly in males at            per day
                      2500, 5000, 10 000, 20 000 or     >5000 mg/kg and in females at >10 000 mg/kg;        (peroxisomal
                      40 000 mg/kg; mean equivalent     testis and epididymal weights lower at 20 000       proliferation);
                      dose levels based on              and 40 000 mg/kg                                    720 mg/kg b.w.
                      consumption and b.w.s:            Haematology:  minimal anaemia in males at           per day based
                      females:  0, 177, 356, 712,       >5000 mg/kg                                         on germinal
                      1413 or 2943 mg/kg b.w.           Clinical chemistry:  hypocholesterolaemia in        epithelial
                      males:  0, 176, 359, 720, 1540    both sexes at 20 000 and 40 000 mg/kg;              atrophy in the
                      or 2964 mg/kg b.w.                hypotriglyceridaemia in all exposed males and       testes and
                                                        females at >10 000 mg/kg; elevations in             histopathological
                                                        alkaline phosphatase activity and bile acid         lesions in the
                                                        concentration in both sexes was considered          liver
                                                        indicative of cholestasis
                                                                                                            NOEL = 177
                                                                                                            mg/kg b.w.
                                                                                                            per day
                                                                                                                                            

    Table 7.  Continued

                                                                                                                                            

    Species                   Protocol                                      Results                         Effect Levels      Reference
                                                                                                                                            

    F-344 rat;                                          Histopathology:                                                        NTP (1995)
    10 per sex                                          Liver:  hepatocellular cytoplasmic alterations
    per group                                           consistent with glycogen depletion in both sexes
                                                        at >10 000 mg/kg; in both sexes at
                                                        40 000 mg/kg, small, fine,
                                                        eosinophilic granules in cytoplasm of
                                                        hepatocytes; ultrastructural examination
                                                        showed the presence of increased numbers
                                                        of peroxisomes, and peroxisomal enzyme
                                                        activity was elevated in liver of both
                                                        sexes at >5000 mg/kg (enzyme activity at
                                                        40 000 mg/kg was 13-fold greater than
                                                        controls for males and 32-fold greater
                                                        than controls for females)
                                                        Testes:  degeneration of germinal epithelium,
                                                        mild to marked focal lesions at 10 000 and
                                                        20 000 mg/kg and marked, diffuse lesion in all
                                                        animals at 40 000 mg/kg; almost complete loss
                                                        of germinal epithelium at 40 000 mg/kg;
                                                        testicular zinc concentrations lower at 20 000
                                                        and 40 000 mg/kg; serum testosterone lower
                                                        than controls at 20 000 and 40 000 mg/kg; at
                                                        20 000 mg/kg, spermatid heads per testis and
                                                        per gram testis, epididymal spermatozoal
                                                        motility, and the number of epididymal
                                                        spermatozoa per gram epididymis were lower than
                                                        controls.
                                                                                                                                            

    Table 7.  Continued

                                                                                                                                            

    Species                   Protocol                                      Results                         Effect Levels      Reference
                                                                                                                                            

    F-344 rat         This study has been designated    10 control and 10 exposed pups per sex were         LOEL for hepatic   NTP (1995)
                      "NTP Study 3"                     examined at weaning; hepatomegaly and               peroxisomal
                                                        markedly increased peroxisomal enzyme               proliferation and
                      Combined perinatal and            activities (19-fold greater than control            hepatomegaly is
                      subchronic exposure. An           values) were observed.                              279 mg/kg b.w. per
                      unspecified number of dams was    Body weight gain:  dose-related decrease            day for males and
                      administered 10 000 mg/kg in      significant in males in all dose groups and         593 mg/kg b.w. per
                      the diet beginning at day one     in females at >20 000 mg/kg                         day for females;
                      of gestation, throughout          Organ weights: hepatomegaly in males at             NOEL is 138 mg/kg
                      gestation, and until weaning.     >5000 mg/kg and in females at >2500 mg/kg;          b.w. per day for
                      Pups had no exposure for four     lower testes weight at >20 000 mg/kg; lower         males and 294 mg/kg
                      weeks post-weaning.  Ten pups     epididymal weight at 20 000 mg/kg                   b.w. per day for
                      per sex per group were then       Haematology: mild anaemia in males at               females.
                      administered 2500, 5000,          >10 000 mg/kg and in females at 40 000 mg/kg
                      10 000, 20 000 or 40 000 mg/kg    Clinical chemistry:  hypocholesterolaemia in
                      in the diet for 13 weeks. There   both sexes at higher concentrations;
                      were two control groups -         hypotriglyceridaemia in females at 20 000
                      animals which had been            and 40 000 mg/kg and in males at >10 000
                      perinatally exposed and those     mg/kg; elevations in alkaline phosphatase
                      with no exposure.                 activity and bile acid concentrations in
                                                        both sexes at 20 000 and 40 000 mg/kg were
                                                        indicative of cholestasis
                                                                                                                                            

    Table 7.  Continued

                                                                                                                                            

    Species                   Protocol                                      Results                         Effect Levels      Reference
                                                                                                                                            

    F-344 rat         Mean equivalent dose levels       Histopathology, liver:  hepatocellular              testicular         NTP (1995)
                      based on consumption and body     cytoplasmic alteration, consistent with             germinal
                      weights:                          glycogen depletion, in both sexes at                epithelium
                      females: 0, 147, 294, 593, 1182   >10 000 mg/kg; small, fine, eosinophilic            degeneration 
                      and 2445 mg/kg b.w. per day       granules in cytoplasm of hepatocytes in             observed at 571
                      males: 0, 138, 279, 571, 1262     males at 40 000 mg/kg; ultrastructural              mg/kg b.w. per
                      and 2495 mg/kg b.w. per day       examination showed the presence of                  day
                                                        increased numbers of peroxisomes at the
                                                        highest dose; peroxisomal enzyme activity
                                                        increased in males at >5000 mg/kg and in
                                                        females at >10 000 mg/kg (at 40 000 mg/kg,
                                                        activities were 20-fold higher than in
                                                        controls) testes:  degeneration of germinal
                                                        epithelium, mild to moderate focal lesion
                                                        at 10 000 and 20 000 mg/kg, and marked,
                                                        diffuse lesion at 40 000 mg/kg; almost
                                                        complete loss of germinal epithelium at
                                                        40 000 mg/kg; testicular zinc concentrations
                                                        lower at 40 000 mg/kg; at 20 000 mg/kg,
                                                        there were fewer spermatid heads per testis
                                                        than unexposed controls and epididymal
                                                        spermatozoal concentrations were less than
                                                        in perinatally exposed controls

    Rat,              ingestion of diets containing     half of the animals at the highest dose died                           Smith (1953)
    Sprague-Dawley,   0.01, 0.05, 0.25 or 1.25% DBP     during the first week
    10 males per      for 1 year
    group                                               haematology:  no abnormalities at 3, 6, 9
                      equivalent doses:  6, 30, 150     months or at necropsy
                      or 750 mg/kg b.w. per day
                                                        no gross or microscopic changes in lung,
                                                        heart, liver, spleen, adrenals, kidneys,
                                                        stomach, small intestine, thyroid or brain.
                                                                                                                                            

    Table 7.  Continued

                                                                                                                                            

    Species                   Protocol                                      Results                         Effect Levels      Reference
                                                                                                                                            

    Rat, Wistar,      ingestion of diet containing 0    15% of exposed rats died                                               Nikonorow
    20 males and 20   or 0.125% DBP for 1 year;                                                                                et al. (1973)
    females per                                         no gross or histological changes in liver,
    group             equivalent dose:  0 or 75         kidney or spleen
                      mg/kg/b.w. per day

    Mouse (ddy,       0.25% or 2.5% in diet for 86      remarkable vacuolar degeneration and                LOAEL = 500        Ota et al. 
    groups of 3       or 90 days; (500 or 5000 mg/kg    necrosis of single cells in the liver, and          mg/kg b.w.         (1973, 1974)
    males and 12      b.w. per day)                     cysts and degeneration of epithelial cells          per day
    females)                                            in the renal tubules in the high-dose group;
                                                        in the low-dose group, histological changes
                                                        were slight in the liver and kidney but
                                                        degeneration of parenchyma was observed

    B6C3F1 mice; 10   administration in diet for 13     Mean body weight and body weight gain: decreased    LOEL = 812         NTP (1994a,b;
    per sex per       weeks (0, 1250, 2500, 5000,       in both sexes in a dose-related manner; decreases   mg/kg b.w.         1995)
    group             10 000 or 20 000 mg/kg);          were significant at >5000 mg/kg                     per day, based
                      mean equivalent dose levels       Relative liver weight:  greater in both sexes at    on decreases
                      based on consumption and body     >5000 mg/kg                                         in body weight
                      weights:                          Haematology:  minimal anaemia was suggested in      in both sexes
                      females: 0, 238, 486, 971,        females at 20 000 mg/kg                             and increases
                      2137, or 4278 mg/kg b.w. per      Histopathology:  hepatocellular cytoplasmic         in relative
                      day                               alterations, consistent with glycogen depletion,    liver weight
                      males: 0, 163, 353, 812, 1601     in males at 10 000 and 20 000 mg/kg and females     (NTP, 1995)
                      or 3689 mg/kg b.w. per day        at 20 000 mg/kg; small, fine, eosinophilic
                                                        granules consistent with peroxisome proliferation   NOEL = 486
                                                        were observed in the cytoplasm of hepatocytes in    mg/kg b.w. per
                                                        both sexes at 20 000 mg/kg.                         day
                                                                                                                                            

    Table 7.  Continued

                                                                                                                                            

    Species                   Protocol                                      Results                         Effect Levels      Reference
                                                                                                                                            

    Inhalation

    Rat (strain,      900 ± 80 mg/m3, 6 h per day       Reduced body weights in exposed animals at the      one dose group     Antonyuk & 
    number and sex    for 35 days                       end of the study (initial 173 g, final 151 g)       only (effects      Aldyreva
    not specified)                                      compared with controls (initial 174 g; final        observed at        (1973)
                                                        223 g).  A decrease in haemoglobin content of       900 mg/m3)
                                                        peripheral blood was observed on day 10 but there
                                                        was a slight increase over control values by day
                                                        30.  A decrease in phagocytic ability of
                                                        peripheral blood neutrophils was also noted. 
                                                        There were statistically significant increases
                                                        in the weight of the liver, lungs, kidneys,
                                                        adrenals and brain, but data were not presented.

    Mouse (strain     range of 20 to 85 mg/m3 for 86    At the end of the study period, pulmonary           could not be       Spasovski
    and sex not       days.  For a further 6 days,      oedema was observed.  No other end-points           determined         (1964)
    specified,        the concentration was increased   were reported.
    groups of 15)     to a range of 170 to 420 mg/m3
                                                                                                                                            

    Table 7.  Continued
                                                                                                                                            

    Species                   Protocol                                      Results                         Effect Levels      Reference
                                                                                                                                            

    Rat (Wistar,      0.5 mg/m3 (0.04 mg/kg) or         A reduced rate of body weight gain was noted        LOEL = 0.5         Kawano
    groups of 11 to   50 mg/m3 (4.4 mg/kg) vapour,      in rats administered the high concentration.        mg/m3              (1980a)
    14 males)         6 h per day except 1 day per      There were no major differences between controls
                      week when exposure was for        and exposed animals with respect to red cells,
                      3 h, 6 days/week for up to 6      platelets, haemoglobin concentration, haematocrit,
                      months.  The levels of DBP        lymphocytes, and neutrophils at 3 months but there
                      were checked throughout the       was a reduction in lymphocyte numbers with an
                      study and varied between 45.0     increase in neutrophil numbers in exposed animals
                      and 59.2 mg/m3 in the high        at 6 months. There were slight increases in
                      concentration and 0.31 and        serum aspartate aminotransferase, alanine
                      0.56 mg/m3 in the low             aminotransferase and alkaline phosphatase levels
                      concentration chambers.           at 1, 3 and 6 months, and in blood glucose level
                                                        at 6 months in animals exposed to both
                                                        concentrations.  At 6 months, serum cholesterol
                                                        had fallen slightly but serum triglyceride levels
                                                        had risen.  The effects in the rats in the
                                                        low-concentration group were similar, but less
                                                        pronounced.  In the high-concentration group, the
                                                        relative organ weights (brain, lung, liver,
                                                        kidney, testes) were normal at the end of the
                                                        first month, but after 6 months the relative
                                                        weights of the brain, lung, kidney and testes had
                                                        all increased.

    Rat (strain       0.095, 0.25 and 1 mg/m3, 24 h     No clinical signs of toxicity were noted during     NOEL = 1 mg/m3     Men'shikova
    unspecified,      per day for 93 days               the study; animals appeared healthy and the rate                       (1971)
    groups of 15                                        of body weight gain for treated animals was
    males)                                              similar to controls.  No abnormalities in
                                                        haemoglobin or red blood cell counts based on
                                                        haematological examinations, which were carried
                                                        out monthly.  White cell counts fell during
                                                        exposure and these did not return to normal in a
                                                        group of animals observed for 6 months after
                                                        termination of exposure.
                                                                                                                                            
        NTP study 1: prenatal/postnatal dose range-finding study in F-344
    rats (described in section 7.5)
    NTP study 2: conventional 13 week study in F-344 rats
    NTP study 3: prenatal/postnatal exposure followed by a 13-week
    exposure with two control groups (one with and one without
    prenatal/postnatal exposure) in F-344 rats
    NTP study 4: continuous breeding protocol in Sprague Dawley rats
    (described in section 7.5)

         No deaths occurred in the 13-week dietary study (NTP study 2
    in rats) in which groups of 10 F-344 rats received 0, 2500, 5000,
    10 000, 20 000 or 40 000 mg/kg diet.  Average equivalent doses
    were 0, 177, 356, 712, 1413 or 2943 mg/kg body weight per day for
    females and 0, 176, 359, 720, 1540 or 2964 mg/kg body weight per
    day for males (NTP, 1995).  The final body weights of males
    receiving approximately > 720 mg/kg body weight per day and
    females receiving approximately  > 1413 mg/kg body weight per
    day were less than those of controls.  No overt hepatic necrosis
    or inflammation was observed at any dose.  Hepatomegaly was
    observed in males exposed to approximately > 359 mg/kg body
    weight per day and females exposed to > 712 mg/kg body weight
    per day.  Testis and epididymal weights were less than those of
    controls in animals exposed to > 1540 mg/kg body weight. 
    Histopathological examination of the liver revealed
    hepatocellular cytoplasmic alterations, consistent with glycogen
    depletion, in both sexes at >10 000 mg/kg diet.  At 40 000
    mg/kg diet, eosinophilic granules were observed in hepatocellular
    cytoplasm. Upon ultrastructural examination, increased numbers of
    peroxisomes were observed, and peroxisomal enzyme activity was
    elevated in the liver of both sexes at > 5000 mg/kg diet. 
    Hepatic peroxisomal enzyme activity at the highest dose was 13-
    and 32- fold greater than controls in males and females,
    respectively.  Examination of the testes revealed degeneration of
    the germinal epithelium, mild to marked focal lesions at
    > 720 mg/kg body weight,  and a marked, diffuse lesion in all
    males at 2964 mg/kg body weight per day.  There was an almost
    complete loss of the germinal epithelium at this dose. 
    Concentrations of testicular zinc and serum testosterone were
    less than those of controls at > 1540 mg/kg body weight per
    day.  At 2400 mg/kg body weight per day, spermatid heads per
    testis and per gram testis, epididymal spermatozoal motility, and
    the number of epididymal spermatozoa per gram epididymis were
    less than that in controls.

         The only target tissues identified, therefore, in this study
    were the liver and testes. The LOEL and NOEL values in this study
    for hepatic peroxisomal proliferation and hepatomegaly were 356
    and 176 mg/kg body weight per day, respectively.  Testicular
    germinal epithelial degeneration was observed at higher doses
    (720 mg/kg body weight per day).

         In a NTP study 3, pregnant F-344 rats were administered 0 or
    10 000 mg/kg in the diet during gestation and lactation, and
    weaned pups were administered the same diets as their dams
    received for an additional 4 weeks until the beginning of the 13-
    week exposure phase (NTP, 1995).  The offspring then received 0,
    2500, 5000, 10 000, 20 000 or 40 000 mg/kg diet (equivalent to
    mean doses of 0, 147, 294, 593, 1182 and 2445 mg/kg body weight
    per day for females and 0, 138, 279, 571, 1262 and 2495 mg/kg
    body weight per day for males) for 13 weeks.

         Ten control and 10 exposed pups of each sex were examined at
    weaning.  Hepatomegaly was observed in exposed pups, and
    peroxisomal enzyme activity was 19-fold greater than in controls. 
    The body weight of prenatally/perinatally exposed pups was less
    than that of controls throughout the 4-week period prior to the
    13-week adult exposures. At the end of the 13-week exposure, 
    statistically significant changes in prenatally/perinatally
    exposed controls versus unexposed controls were limited to
    increased relative testis weight and lower final body weight in
    males.  The Task Group felt that the appropriate comparison was
    to the pretreated control groups.

         At the end of the 13-week exposure, body weights of males in
    all exposed groups and of females at > 593 mg/kg body weight
    per day were less than those in unexposed controls.  No overt
    hepatic necrosis or inflammation was observed at any dose.  In
    adult rats, hepatomegaly was observed in males at > 279 mg/kg
    body weight per day and in females at > 593 mg/kg body weight
    per day. There was hepatocellular cytoplasmic alteration 
    consistent with glycogen depletion in both sexes at >
    10 000 mg/kg diet.  Marked elevations of peroxisomal enzyme
    activity were detected in males receiving > 279 mg/kg body
    weight per day and in females receiving > 593 mg/kg body
    weight per day.

         At > 1262 mg/kg body weight per day, testis weight was
    lower than that in controls.  There was mild to moderate
    degeneration of the germinal epithelium at > 571 mg/kg body
    weight per day and marked diffuse germinal epithelial
    degeneration at 2495 mg/kg body weight per day, at which dose an
    almost complete loss of the germinal epithelium resulted.  At the
    highest dose, testicular zinc concentration was reduced, there
    were fewer spermatid heads per testis than in unexposed controls,
    and epididymal spermatozoal concentration was less than that in
    the prenatally perinatally exposed controls.

         The only target tissues identified, therefore, in this study
    were the liver and testes. The LOEL and NOEL values in this study
    for hepatic peroxisomal proliferation and hepatomegaly were 279
    and 138 mg/kg body weight per day in males and 593 and 294 mg/kg
    body weight per day in females.  Testicular germinal epithelial
    degeneration was observed at higher doses (571 mg/kg body weight
    per day).

         In other studies, male Wistar rats were administered 5% DBP
    in the diet (equivalent to 2500 mg/kg body weight per day) for 35
    to 45 days (Murakami et al., 1986b) or 0.5 or 5% (equivalent to
    250 or 2500 mg/kg body weight per day) for 34 to 36 days
    (Murakami et al., 1986a).  In the rats ingesting the diet
    containing 5% DBP, there was growth depression, liver
    enlargement, testicular atrophy, decreased activities of
    succinate and pyruvate dehydrogenase in liver mitochrondria and
    abnormal changes in biochemical tests of serum and in
    histological examinations of the liver and testes.  Hepatic
    lesions (including necrosis and atrophy) were also observed in
    rats fed the diet containing 0.5% DBP, although these lesions
    were less severe than those reported in the high-dose group. 
    There were changes in hepatocellular ultrastructure in rats
    exposed to DBP, which were related to increases in peroxisomes,
    lysosomes and mitochondria.  A NOEL could not be established on
    the basis of this study; the LOAEL was 250 mg/kg body weight per
    day, based on liver pathology (Murakami et al., 1986b).

         Additional mechanistic studies into DBP-related effects on
    hepatic cell proliferation, peroxisomal enzyme activities,
    clinical chemistry, haematology and gene expression are being
    undertaken in a 90-day feed study by the National Toxicology
    Program, but published reports are not yet available (personal
    communication by R.R. Maronpot, National Institute of
    Environmental Health Sciences, to the IPCS, 1995).

         Little information on repeated dose toxicity in rats
    following ingestion for periods longer than 3 months has been
    identified.  In an early study (Smith, 1953), groups of 10 male
    Sprague-Dawley rats ingested diets containing 0.01, 0.05, 0.25 or
    1.25% DBP (equivalent to 6, 30, 150 or 750 mg/kg body weight per
    day) for 1 year.  There were no effects on growth, but half of
    the exposed animals administered the highest dietary level
    (750 mg/kg body weight per day) died during the first week of the
    study.  No abnormalities were seen during examination of
    haematological parameters at 3, 6 and 9 months, and at necropsy
    there were no abnormal gross or microscopic findings in any of
    the organs examined, which included the lung, heart, liver,
    spleen, adrenals, kidneys, stomach, small intestine, thyroid and
    the brain.

         In another 12-month dietary study (Nikonorow et al., 1973),
    groups of 20 male and 20 female Wistar rats ingested a diet
    containing 0.125% of DBP (equivalent to 75 mg/kg body weight per
    day).  There were marked differences in food intake in exposed
    animals, compared to controls, and 15% of the exposed rats died. 
    No alterations in the liver, kidneys or spleen were seen during
    gross and histological examination.

         In an inhalation study, effects at lowest levels were
    observed by Kawano (1980b).  In this investigation, rats were
    exposed to 0.5 or 50 mg/m3 of DBP mist for 6 h/day, 5 days a
    week, for up to 6 months (except for one day/week when rats were
    exposed for only 3 h).  There was a reduction in the rate of body
    weight gain and increases in the relative weights of the brain
    and lung.  An increase in the percentage of neutrophils was
    observed in both exposed groups.  High levels of urea nitrogen
    and low levels of cholesterol and triglyceride in the serum of
    rats exposed to the high concentration were observed, indicating
    hypolipidaemic activity of DBP.  Similar, but less pronounced,
    effects were observed in the group exposed to the low
    concentration.  In other identified studies, no effects were
    observed following exposure for 93 days to 1 mg/m3 (Men'shikova,
    1971), whereas effects on body weight gain, organ weights and
    haematological parameters were observed at a high concentration
    (900 mg/m3) following exposure for 35 days (Antonyuk &
    Aldyreva, 1973).

         In small groups of male and female ddY mice given a diet
    containing 500 or 5000 mg/kg body weight per day for 3 months,
    there were marked lesions in the liver and kidney (Ota et al.,
    1973, 1974).  In the high-dose group, there was remarkable
    vacuolar degeneration and necrosis of single cells in the liver,
    and cysts and degeneration of tubular epithelial cells in the
    kidney.  In the low-dose group, histological changes were slight
    in the liver and kidney but degeneration of the parenchyma was
    observed.  The LOAEL  was considered to be 500 mg/kg body weight
    per day on the basis of histological changes in the liver and
    kidneys.

         B6C3F1 mice received dietary concentrations of 0, 1250,
    2500, 5000, 10 000 or 20 000 mg/kg diet for 13 weeks (equivalent
    to mean doses of 0, 238, 486, 971, 2137, or 4278 mg/kg body
    weight per day in females and 0, 163, 353, 812, 1601 or
    3689 mg/kg body weight per day in males (NTP, 1995).  Body weight
    and body weight gain were significantly decreased and relative
    liver  weight was significantly increased in both sexes at
    > 5000 mg/kg diet.  No overt hepatic necrosis was observed. 
    Histopathological examination revealed hepatocellular cytoplasmic
    alterations, consistent with glycogen depletion, in males at
    >1601 mg/kg body weight per day and in females at 4278 mg/kg
    body weight per day.  Eosinophilic granules, consistent with
    peroxisomal proliferation, were observed in the cytoplasm of
    hepatocytes in both sexes at 20 000 mg/kg diet. The LOEL in this
    study was 812 mg/kg body weight per day, based on decreases in
    body weight in both sexes and increases in relative liver weight
    (NOEL = 353 mg/kg body weight per day).  In contrast to rats,
    there were no histological alterations in the testes of mice
    exposed to any dose of DBP.

         In summary, the effects in rats following ingestion of DBP
    for periods of up to 13 weeks include reduced rate of weight gain
    at doses of >250 mg/kg body weight per day (Radeva & Dinoyeva,
    1966; Murakami et al., 1986a; NTP, 1995).  Increase in relative
    liver weights have been observed at doses of >120 mg/kg body
    weight per day (Nikonorow et al., 1973; Murakami et al., 1986a;
    1986b; NTP, 1995).  Peroxisomal proliferation as determined by
    increased peroxisomal enzyme activity has been observed at doses
    of >279 mg/kg body weight per day (NTP, 1995).  Although in
    Fischer rats, no overt pathological effects on the liver were
    observed (NTP, 1995), necrotic hepatic changes in Wistar rats
    have been reported at doses of >250 mg/kg body weight per day
    (Murakami et al., 1986a).  Effects on the testes of male rats
    have been observed at doses of >571 mg/kg body weight per day
    (NTP, 1995).

         In summary, histopathological lesions in the kidney and
    liver were observed in mice in a limited study at DBP doses of
    > 500 mg/kg body weight per day for 3 months (Ota et al.,
    1973, 1974).  Effects on body and organ weights and histological
    alterations in the liver were also reported  at higher doses in a
    subchronic bioassay on mice (NTP, 1995) for which the NOEL was
    353 mg/kg body weight per day, which was the lower of the two
    NOELs in males and females (NTP, 1995).

         In rats given a diet containing 5% MBP (the principal
    metabolite of DBP), growth depression, liver enlargement,
    testicular atrophy, decreased activities of succinate and
    pyruvate dehydrogenase in liver mitochrondria, biochemical
    effects on serum and histopathological effects on the liver and
    testes were noted (Murakami et al., 1986a).  Hepatic necrosis was
    also  observed in rats fed a diet containing 0.5% MBP (250 mg/kg
    body weight per day), although it was less severe than in the
    rats administered the higher concentration.  The changes in
    hepatocellular ultrastructure were more prominent in rats exposed
    to DBP than in those administered MBP.

    7.4  Irritation and sensitization

         DBP appears to have little potential to irritate skin.  Very
    slight skin irritation, but no skin sensitization, was seen when
    over 4 mg/kg body weight per day was applied to the skin of
    rabbits for 90 days (Lehman, 1955).  The ability of DBP to induce
    an inflammatory response based on extravasation of trypan blue
    has also been examined in rabbits (Calley et al., 1966).  An
    inflammatory response was observed but no details were given as
    to whether there was any clinical irritancy.

         No irritation was noted after DBP was instilled into the
    eyes of rabbits examined at intervals up to 48 h after
    application (concentration or amount not specified) (Lawrence
    et al., 1975).

    7.5  Reproductive and developmental toxicity

    7.5.1  Reproductive effects

    7.5.1.1  Testicular effects

         Studies on the reproductive effects of DBP are summarized in
    Table 8.  Repeated oral exposure to concentrations of DBP for 4
    to 90 days (250 to 2600 mg/kg body weight per day) affects the
    reproductive system of male rodents.  However, there are
    considerable interspecies differences in response.  Observed
    effects in the available studies, most of which only used one
    dose level (generally in the range 1200 to 2400 mg/kg body weight
    per day), included marked reductions in the weights of the testes
    and accessory sex glands, decreased numbers of spermatocytes,
    degeneration of the seminiferous tubules of the testes, a
    reduction in testicular zinc and iron levels and serum
    testosterone levels, an increase in testosterone levels in the
    testes, sloughing of germ cells, decreased activity of succinate
    dehydrogenase in Sertoli cells, and an increase in urinary zinc
    excretion at doses of >250 mg/kg body weight per day (Cater et
    al., 1977; Gray & Butterworth, 1980; Oishi & Hiraga, 1980a,
    1980b; Gray et al., 1982; Ikemoto et al., 1983; Fukuoka et al.,
    1989, 1990; Zhou et al., 1990; Srivastava et al., 1990a,b; Lake
    et al., 1991; Fukuoka et al., 1993; NTP, 1995).

         The lowest reported effect levels in sufficiently well-
    documented studies were those in a multi-dose investigation in
    which DBP in groundnut oil (250, 500 or 1000 mg/kg body weight
    per day) was administered to young male rats by gavage for 15
    days (Srivastava et al., 1990a,b).  A significant decrease in the
    weight of the testes was observed at 500 and 1000 mg/kg body
    weight per day.  At these two doses, histopathological
    examination revealed marked degeneration of the seminiferous
    tubules.  In all exposed groups, the activities of testicular
    enzymes associated with post-meiotic spermatogenic cells, such as
    sorbitol dehydrogenase and acid phosphatase, were decreased
    significantly (P < 0.05), while that of testicular specific
    lactate dehydrogenase was significantly increased, coincident
    with degeneration of spermatogenic cells.  The activities of
    enzymes associated with pre-meiotic spermatogenic cells, Sertoli
    cells or interstitial cells, and of ß-glucuronidase,
    gamma-glutamyl-transpeptidase and glucose-6-phosphate
    dehydrogenase were significantly increased (P < 0.05). 
    Therefore, the LOEL was 250 mg/kg body weight per day based on
    enzyme changes in the testes; this value was also the NOAEL for
    testicular weight and histopathological changes.

        Table 8.  Reproductive effects of DBP

                                                                                                                                           

    Species                  Protocol                                Results                            Effect Levels        Reference
                                                                                                                                           

    Rat               500, 1000 or 2000 mg/kg       In the first study, decreases in testicular         LOEL = 500 mg/kg     Cater et al.
    (Sprague-Dawley,  b.w. per day DBP by           weight at the two highest doses (p<0.01 and         b.w. per day in      (1977)
    groups of 6       gavage in corn oil daily      p<0.001) after 4 days; the weight decreased         one experiment.
    males, 3 to 4     for 4 or 6 days in one        further after 8 days at all 3 doses (p<0.05         One dose group
    weeks old)        experiment.  2000 mg/kg       at 500 mg/kg b.w. per day and p<0.001 at the        only in the other
                      b.w. per day by gavage        two highest doses).  In the second study, body      experiment
                      in corn oil daily for         weight gain also declined but the change was not    (effects observed
                      periods of up to 14 days      significant.  Diminution of both spermatocyte       at 2000 mg/kg
                      in a second experiment.       and spermatogonia counts upon histological          b.w. per day)
                      For urinary zinc              examination of testes after 4 days of exposure
                      measurements, zinc-65         to 2000 mg/kg b.w. per day.  At 2000 mg/kg b.w.
                      chloride was administered     per day, urinary zinc excretion was increased
                      and the zinc-65 content       by 34 to 43% over the first 4 days and then
                      was estimated by              returned to normal.  In the testes, the turnover
                      radioactive counting.         rate of zinc was increased, the half-life was
                                                    reduced from 14 to 5 days and the zinc levels
                                                    were significantly reduced in the testes.
                                                    There was no change in zinc half-life or content
                                                    in liver or kidneys.  Specific activity of
                                                    testicular alcohol dehydrogenase (or
                                                    zinc-dependent enzyme) decreased to 20 to 40%
                                                    of control after 5 days of exposure.  A 3-day
                                                    pretreatment with 2000 mg/kg b.w. per day
                                                    caused a 25% decrease (p<0.001) in testicular
                                                    zinc uptake in vivo.  Simultaneous ingestion of
                                                    zinc sulfate resulted in no testicular atrophy,
                                                    as measured by relative testicular weight (no
                                                    histopathological examinations).
                                                                                                                                           

    Table 8.  Continued

                                                                                                                                           

    Species                  Protocol                                Results                            Effect Levels        Reference
                                                                                                                                           

    Rat               2% in the diet                Mean body weights were slightly but not             One dose group       Oishi &
    (JCL:Wistar,      (equivalent to 2060           significantly lower than that of the controls.      only (effects        Hiraga
    groups of 10      mg/kg b.w. per day)           Absolute and relative testicular weights were       observed at 2060     (1980a)
    males, 5 weeks    for 1 week.                   significantly decreased, and both absolute and      mg/kg b.w. per
    old)                                            relative liver weights were significantly           day)
                                                    increased.  Histological examination of the
                                                    testes revealed a decrease in both spermatocyte
                                                    and spermatogonia counts.  Zinc concentrations
                                                    in the testes and the liver were significantly
                                                    decreased.  Testosterone concentration in the
                                                    testes was significantly increased.

    Rat (Wistar,      2000 mg/kg b.w. per day       Decreases in the relative weight of the testes,     One dose group       Gray & 
    groups of 10      by oral intubation in corn    prostate and seminal vesicles were reported         only (effects        Butterworth
    males, 4 weeks    oil daily for 10 days.        (data not presented).  Testicular atrophy was       observed at 2000     (1980)
    old                                             observed.                                           mg/kg b.w. per day)

    Rat               2000 mg/kg b.w. per day       There was no change in body weight.  Weight of      One dose group       Gray et al.
    (Sprague-Dawley,  by oral intubation in         the testes was reduced to 45% of control            only (effects        (1982)
    groups of 6       corn oil daily for 9 days.    (p<0.001), and >90% tubular atrophy was seen        observed at 2000
    males, 4 to 6                                   in all animals upon histological examination.       mg/kg b.w. per day)
    weeks old)
                                                                                                                                           

    Table 8.  Continued

                                                                                                                                           

    Species                  Protocol                                Results                            Effect Levels        Reference
                                                                                                                                           

    Rat (Wistar,      2400 mg/kg b.w. per day       Severe testicular atrophy was evident; the weight   One dose group       Ikemoto et
    groups of 5       administered by gavage        of the testes was 0.59 g compared with 0.97 g in    only (effects        al. (1983)
    rats, 5-week      (vehicle unspecified)         controls.  Relative organ weights were not          observed at 2400     (English
    old males)        daily for 7 days.             presented and body weights were presented           mg/kg b.w. per day)  abstract)
                      Animals were killed           graphically, only.  Based on these graphs, the
                      on the day after              relative weight of the testes was approximately
                      administration terminated.    0.29% compared with 0.48% for controls.  Upon
                                                    microscopic examination, there was an almost
                                                    complete absence of germ cells in the
                                                    seminiferous tubules, with enlargement and
                                                    vacuolation of the Sertoli cells.  These cells
                                                    had increased numbers of lipid droplets.  The
                                                    Leydig cells also appeared atrophied.  The gross
                                                    and microscopic changes were accompanied by a
                                                    decrease in the serum testosterone levels to 82%
                                                    of control values.

    Rat (Wistar,      2400 mg/kg b.w. per day       Decreases in testicular fructose and glucose        One dose group       Fukuoka et
    28 adult males)   administered by gavage        levels and a sloughing of the germ cells on the     only (effects        al. (1989)
                      (neat DBP) daily for 7        first day of exposure.  On day 2, more severe       observed at 2400
                      days after acclimatization    sloughing, accompanied by decreases in testicular   mg/kg b.w. per day)
                      for 1 week.  Groups of 6      iron and zinc levels and increases in the level     
                      exposed rats were killed at   of inositol and cholesterol.  The sloughing was     
                      24, 48, 120 or 168 h and 2    followed by atrophy, accompanied by dissociation
                      were killed at 72 and 96 h.   of the germ cells from the Sertoli cells and
                                                    reduction of triglycerides, cholesterol and
                                                    phospholipids containing choline and ethanolamine
                                                    residues in the testis.
                                                                                                                                           

    Table 8.  Continued
                                                                                                                                           

    Species                  Protocol                                Results                            Effect Levels        Reference
                                                                                                                                           

    Rat (Wistar,      Single oral dose of 2400      Based on histological examination, DBP caused       One dose group       Fukuoka et
    27 adult males)   mg/kg b.w. per day            sloughing of the germ cells at 6 h.  On days 1      only (effects        al. (1990)
                      administered by gavage        and 2, there was more severe sloughing, followed    observed at 2400
                      (neat DBP) after              by atrophy and the dissociation of the germ cells   mg/kg b.w. per day)
                      acclimatization for 1 week.   from the Sertoli cells and the spermatogonia.
                      Groups of 3 exposed rats      Biochemically, there was an elevation of
                      were killed at 3, 6, 12, 24,  gamma-glutamyl transferase, a decrease in sorbitol
                      48, 72, 96, 120 and 168 h.    levels at 3 h up to the 7th day and a decrease in
                                                    the activity of aldose reductase at 6 h in the
                                                    testes of treated rats.  This was followed by
                                                    decreases in fructose levels and increases in the
                                                    activity of lactate dehydrogenase (LDH) and in
                                                    lactate levels at 12 h, and decreases in the
                                                    activities of sorbitol dehydrogenase and succinate
                                                    dehydrogenase on day 2.  LDH isoenzymes 4 and 5
                                                    increased at 6 h prior to the increase in lactate
                                                    levels.  The data are consistent with DBP-induced
                                                    testicular toxicity being associated with a
                                                    disturbance of the activity of the enzymes that
                                                    are linked with Sertoli cell function and
                                                    replication and germ cell maturation.

    Rat (Wistar       All treated groups            Mono-n-butyl phthalate (MBP) (metabolite of DBP)    One dose group       Zhou et
    adult male)       administered neat DBP as      was transported through the blood-tubular barrier   only (effects        al. (1990)
                      a single oral dose of 2400    onto the seminiferous lumen; it was incorporated    observed at 2400
                      mg/kg b.w.  Control rats      into the lumen at a maximum rate between 1 and 3 h  mg/kg b.w.)
                      administered 0.9% saline      after dosing.  MBP caused decreases in the
                                                    activities of succinate dehydrogenase in the
    Rat (18)          Experiment A                  Sertoli cells and sorbitol dehydrogenase in the
                      Study of transportation of    germ cells, an increase in the activity of lactate
                      DBP from interstitial cell    dehydrogenase in the germ cells and in the
                      fraction to Sertoli cells     seminiferous lumen and a decrease in testicular
                      and germ cells.  Three rats   iron levels.
                      sacrificed at 1, 3, 6, 12,
                      24 and 48 h.
                                                                                                                                           

    Table 8.  Continued

                                                                                                                                           

    Species                  Protocol                                Results                            Effect Levels        Reference
                                                                                                                                           

    Rat (15 exposed  Experiment B                                                                                            Zhou et
    and 15 control)  Determination of enzyme                                                                                 al. (1990)
                     activities in separated
                     cell fractions.  Three rats
                     in each group sacrificed at
                     3, 6, 12, 24 and 48 h

    Rat (27 exposed  Experiment C
    and 15 control)  Measurement of metal ions in
                     testes.  Three exposed rats
                     sacrificed at 3, 6, 12, 24,
                     48, 72, 96, 120 and 168 h.
                     Three control rats sacrificed
                     at 3, 24, 46, 96 and 168 h.

    Rat (Wistar,     Single oral dose, 2400         At 6 h:  sloughing of germ cells; decrease in       One dose group       Fukuota et
    adult male)      mg/kg b.w. Control rats        activity of succinate dehydrogenase in the          only (effects        al. (1993)
                     received 0.9% saline.          Sertoli cells and in the Sertoli-germ connection;   observed at 2400
                     Serial sacrifice of 3 rats     increase in activity of lactate dehydrogenase in    mg/kg b.w.)
                     at each of 1, 3 and 6 h.       germ cells.  Increases in transferrin
                                                    concentrations in Sertoli cells, Sertoli-germ
                                                    connection, epididymus-ductus deferens and liver.
                                                    Decrease in transferrin in seminal vesicle.
                                                    Decrease in ferritin in seminiferous lumen.
                                                    Increase in flavin adenine dinucleotide level in
                                                    interstitial cells.
                                                                                                                                           

    Table 8.  Continued

                                                                                                                                           

    Species                  Protocol                                Results                            Effect Levels        Reference
                                                                                                                                           

    Rat (Wistar       0, 250, 500 or 1000 mg/kg     Significant decrease in testicular weight at 500    LOAEL = 250 mg/kg    Srivastava
    albino, groups    b.w. per day by gavage in     and 1000 mg/kg b.w. per day.  Histopathological     b.w. per day         et al.
    of 6 males,       ground nut oil daily for      examination revealed marked degeneration of                              (1990a,b)
    5 weeks old)      15 days.                      seminiferous tubules at these doses.  In all
                                                    exposed groups, the activities of testicular
                                                    enzymes associated with post-meiotic spermatogenic
                                                    cells, such as sorbitol dehydrogenase and acid
                                                    phosphatase, were decreased significantly, while
                                                    that of lactate dehydrogenase was significantly
                                                    increased, coincident with degeneration of
                                                    spermatogenic cells.  The activities of enzymes
                                                    associated with premeiotic spermatogenic cells,
                                                    Sertoli cells or interstitial cells,
                                                    œ-glucuronidase, gamma-glutamyl transpeptidase and
                                                    glucose-6-phosphate dehydrogenase were also
                                                    significantly increased in all exposed groups.

    Rat (F-344,       0, 0.05, 0.1, 0.5, 1.0 or     Based on histological changes and data on organ     NOEL = 515 mg/kg     Lake et al.
    groups of 5       2.5% in the diet for 28       weights, the authors concluded that the NOEL for    b.w. per day         (1991)
    males, 6 weeks    days (conversions to dose     testicular atrophy was 515 mg/kg b.w. per day.                           (abstract)
    old)              on a body weight basis not    No other information was reported.
                      available since food
                      consumption was determined
                      but not fully reported)

    Rats (males;      Oral administration by        Some regeneration of seminiferous tubules 2 weeks   One dose group       Tanino et
    strain and        gavage (vehicle               after discontinuation of the administration.        only (effects        al. (1987)
    number of         unspecified) of 2400          Active spermatogenesis in almost all tubules        observed at 2400
    animals           mg/kg b.w. per day DBP        though vacuolation of germinal epithelium and       mg/kg b.w. per day)
    unspecified)      daily for 7 days.             decreased number of sperm were still evident 3
                                                    weeks after exposure was terminated.
                                                                                                                                           

    Table 8.  Continued

                                                                                                                                           

    Species                  Protocol                                Results                            Effect Levels        Reference
                                                                                                                                           

    Rat (males and    500 or 1000 mg/kg b.w. per    In the first experiment, there were no effects      In the first         Gray et
    females, strain   day from 20 to 55 days of     on the female reproductive system while male rats   experiment: LOAEL    al. (1983)
    and number        age in one experiment and     were severely affected at both doses.  Exposed      = 500 mg/kg b.w.     (abstract)
    unspecified)      250 or 500 mg/kg b.w. per     rats had smaller testes and seminal vesicles and    per day (males);
                      day from 20 to 75 days of     no sperm in the vas deferens.  In the second        NOEL = 1000 mg/kg
                      age in a second experiment    experiment, rats were unaffected at 250 mg/kg b.w.  b.w. per day
                      (nature, vehicle and          per day but half of the pairs did not breed in the  (females)
                      pattern of administration     high-dose group.  No other information was          In the second
                      unspecified)                  reported.                                           experiment:
                                                                                                        NOEL=250 mg/kg
                                                                                                        b.w. per day

    Rat (strain       0.52 g/kg b.w. per day by     There were no effects on conception rate or litter  One dose group       Bornmann & 
    unspecified,      gavage daily (vehicle         sizes in exposed animals when compared with         only (no effects     Loeser (1956)
    groups of 8       unspecified) for 6 weeks      controls.  Neonatal growth rates in the F1 were     observed at 520
    females)          and then mated with           also comparable with those of control animals.      mg/kg b.w. per day)
                      untreated males to produce    The weights of endocrine organs in this generation
                      an F1 generation.             were within normal limits and the onset of estrus
                      Interbreeding of untreated    was similar to that in control rats.  There were
                      F1 rats produced an F2        also no abnormalities in the F2 and F3 generations.
                      generation.  An F3
                      generation was similarly
                      produced.
                                                                                                                                           

    Table 8.  Continued

                                                                                                                                           

    Species                  Protocol                                Results                            Effect Levels        Reference
                                                                                                                                           

    Rat (CD           This has been designated      the final dose levels selected were 0.1, 0.5 and    NTP concluded that   NTP (1991,
    Sprague-Dawley)   "NTP Study 4" Continuous      1.0%, on the basis of clinical signs, body weight   these effects        1995); Wine
                      breeding protocol which       and food consumption.                               document             et al. (1997)
                      included cross-over mating                                                        reproductive and
                      and offspring assessment                                                          developmental
                      phases. Preceded by a                                                             toxicity of DBP
                      range-finding study                                                               in F0 rats at all
                      ("Task 1") [5 doses (0.1,                                                         dose levels and
                      0.5, 1.0, 1.5 and 2.0% in                                                         more severe
                      diet) and control; 8                                                              toxicity to F1
                      rats/sex/group]                                                                   offspring.
                                                                                                        NOAEL not
                                                                                                        identified
                                                                                                        LOAEL =
                                                                                                        approximately 66
                                                                                                        mg/kg b.w. per day
                                                                                                                                           

    Table 8.  Continued

                                                                                                                                           

    Species                  Protocol                                Results                            Effect Levels        Reference
                                                                                                                                           

    Rat (CD           "Task 2", continuous          All control and exposed pairs were fertile.         No NOAEL             NTP (1991,
    Sprague-Dawley)   breeding phase. Based upon    Dose-related decrease in number of live pups per    established for      1995); Wine
                      food consumption data, the    litter (significant at all doses); absolute and     decreased litter     et al. (1997)
                      authors estimated that the    adjusted live pup weight significantly decreased    size, LOAEL = 66
                      intakes for the exposed       in mid- and high-dose groups.  Dam weights at       mg/kg b.w. per day.
                      groups were 52, 256 and 509   delivery were significantly decreased at each       NOAEL for pup body
                      mg/kg b.w. for males and      litter in the high dose-group.                      weight 66 mg/kg1
                      80, 385 and 794 mg/kg b.w.    In the mid-dose group, reduction in pup body        b.w. per day.
                      for females, giving average   weight was not accompanied by reduced dam body      NOAEL for fertility
                      intakes of 66, 320 and 651    weight.                                             651 mg/kg b.w. per
                      mg/kg b.w. per day. 40                                                            day.
                      breeding pairs as controls
                      and 3 dose groups of 20
                      pairs each. Animals housed
                      as breeding pairs for 112
                      days.  End-points were
                      clinical signs, parental
                      body weight, feed
                      consumption, fertility
                      (number producing a
                      litter/number of breeding
                      pairs), numbers of litters
                      per pair, number of live
                      pups per litter, proportion
                      of live pups, sex ratio of
                      live pups, body weight of
                      pups.  The last litter born
                      during "Task 2" was reared
                      for "Task 4". litters
                                                                                                                                           

    Table 8.  Continued

                                                                                                                                           

    Species                  Protocol                                Results                            Effect Levels        Reference
                                                                                                                                           

    Rat (CD           "Task 3"; since an adverse    No overall difference with respect to mating,       One dose group       NTP (1991,
    Sprague-Dawley)   effect on reproduction was    pregnancy or fertility indices.  Live pup weight    only.  Effects       1995); Wine
                      detected during Task 2, a     was adversely affected in Group C, which suggested  seen in treated      et al. (1997)
                      1-week cross-over mating      that DBP was a reproductive toxicant in females.    females at 665
                      trial was performed to        However the authors noted that it was not possible  mg/kg b.w. per day.
                      determine the affected sex.   to distinguish between an effect on dam body
                                                    weight and direct toxicity to the fetus.

                      Consisted of 3 groups of 20   Significant increase in organ to body weight
                      pairs each:                   ratios for liver and kidneys in males and F0
                      -  control males × control    females.  No effect upon sperm concentration,
                      females (Group A)             motility, percent abnormal forms or testicular
                      -  high-dose males ×          spermatid head count.  No apparent effects upon
                      control females (Group B)     estrual cyclicity or average estrous cycle length.
                      -  control males ×
                      high-dose females (Group C)
                      End-points as for Task 2,
                      with addition of checking
                      for presence of vaginal
                      copulatory plug or sperm.

                      Estimated average daily
                      intakes of exposed animals
                      were 410 and 665 mg/kg b.w.
                      for males and females
                      respectively.
                                                                                                                                           

    Table 8.  Continued

                                                                                                                                           

    Species                  Protocol                                Results                            Effect Levels        Reference
                                                                                                                                           

    Rat (CD           "Task 4"; the last litter     In the F1 high-dose group, body weight was          No NOAEL             NTP (1991,
    Sprague-Dawley)   born following the            significantly lower at weaning and at necropsy      established for      1995); Wine
                      continuous breeding phase     for both sexes.                                     reduced pup weight.  et al. (1997)
                      ("Task 2") was reared by      Mating, pregnancy and fertility indices were        LOEL = 66 mg/kg
                      the dam until weaning, at     significantly lower in the high-dose group (1       b.w. per day.
                      which time the F1 animals     litter born versus 19 in control group).            No NOAEL
                      were exposed similarly to     Absolute and ajusted live F2 pup weights were       established for
                      the parents until 13 weeks    significantly lower in all exposed groups.          testis tubule
                      of age.  Estimated average    In high-dose males, significant decrease in the     degeneration.
                      daily intakes for exposed     absolute weight and relative weight of prostate,    LOAEL = 322 mg/kg
                      groups were 50, 247 and 498   right testis and seminal vesicles; significant      b.w. per day (low
                      mg/kg b.w. for males and 83,  increase in liver and kidney weight.                dose group not
                      397 and 828 mg/kg b.w. for    In high-dose males, adverse effect upon epididymal  examined)
                      females. At sexual maturity,  sperm count and concentration and testicular
                      groups of 20 males and 20     spermatid head count and concentration.  No
                      females from the same         apparent effects upon estrual cyclicity or average
                      treatment groups cohabited    estrous cycle length.
                      for 7 days, then housed       Epididymides absent or poorly developed in 5/10
                      singly until delivery.        high-dose, 0/10 mid-dose males.
                      End-points same as for        Histological examination of control, mid- and
                      "Task 2", followed by         high-dose groups showed testicular lesions
                      necropsy.                     consisting of degeneration of seminiferous tubules
                                                    (8/10 in high-dose and 3/10 in mid-dose group),
                                                    interstitial cell hyperplasia (7/10 in high-dose
                                                    group) and underdeveloped or defective
                                                    epididymides (5/10 in high dose group).
                                                                                                                                           

    Table 8.  Continued

                                                                                                                                           

    Species                  Protocol                                Results                            Effect Levels        Reference
                                                                                                                                           

    Rat (F-344,       DBP administered in diet to   Weight gain:  decreased in dams at 20 000 mg/kg     10 000 mg/kg diet    NTP (1995)
    number            dams during gestation and     during gestation and in dams at 10 000 mg/kg        was recommended
    unspecified)      lactation and to pups         during lactation.  Mean b.w. of pups reduced        as the maximum
                      postweaning for 4 weeks, at   during lactation and at end of 4 weeks of dietary   perinatal exposure
                      concentrations of 0, 1250,    exposure.                                           concentration for
                      2500, 5000, 7500, 10 000 or   Gestation index: (number of live pups per           male and female
                      20 000 mg/kg diet.            breeding female) was significantly lower in the     rats.
                                                    20 000 mg/kg group [pup mortality in this group
                                                    was 100% by day 1 of lactation]; pup survival was
                                                    89% or more in all other treatment groups.
                                                    Organ weight:  increased relative liver weight in
                                                    all exposed males and in females at >2500 mg/kg.
                                                    Histopathology:  moderate hypospermia of the
                                                    epididymis in all males at 7500 and 10 000 mg/kg;
                                                    mild hypospermia in 2 out of 10 males at 5000
                                                    mg/kg; no degeneration of germinal epithelium was
                                                    detected.

    Mouse (B6C3F1,    Dietary concentrations of 0,  Only 5 dams at 10 000 mg/kg delivered live pups,    Developmental        NTP (1995)
    pregnant, 20      1250, 2500, 5000, 7500,       and none at 20 000 mg/kg.  Only 1 pup at 10 000     toxicity and pup
    per group)        10 000 or 20 000 mg/kg        mg/kg survived past lactation day one; number of    mortality were
                      during gestation and          live pups per litter at 7500 mg/kg remained low     suggested at
                      lactation; pups were weaned   throughout lactation; no deaths occurred in pups    concentrations as
                      onto same diet as dams and    after weaning.                                      low as 7500 mg/kg
                      exposed for an additional     Postweaning and final body weights of males at      and 5000 mg/kg was
                      4 weeks                       >2500 mg/kg were significantly less than controls.  considered to be
                                                    Organ weights:  absolute liver weight of males at   the maximum
                                                    7500 mg/kg was greater than controls.               perinatal exposure
                                                    The one surviving male pup at 10 000 mg/kg had      concentration.
                                                    cytoplasmic alteration in liver, consistent with
                                                    peroxisome proliferation.
                                                    Developmental toxicity and fetal and pup mortality
                                                    were suggested as low as 7500 mg/kg.
                                                                                                                                           

    Table 8.  Continued

                                                                                                                                           

    Species                  Protocol                                Results                            Effect Levels        Reference
                                                                                                                                           

    Mouse (ICR,       2% in the diet (equivalent    Food consumption was affected, though relevant      One dose group       Oishi &
    groups of 10      to 2400 mg/kg b.w. per day)   data were not presented and body weight gain was    only (effects        Hiraga (1980b)
    males)            for 1 week.                   significantly decreased.  The relative weights of   observed at 2400
                                                    the testes and liver were significantly             mg/kg b.w. per day)
                                                    increased, whereas the relative weight of the
                                                    kidney was significantly reduced.  The zinc
                                                    concentration in the testes and liver was reduced
                                                    to 81 ± 3.14% and 88 ± 3.14% of the control,
                                                    respectively (p<0.05), but not in the kidney.
                                                    The concentration of testosterone in the testes
                                                    was unaltered and there was no testicular atrophy.

    Mouse (T O        2000 mg/kg b.w. per day by    There was a significant depression of the weight    One dose group       Gray et 
    strain, groups    oral intubation in corn oil   of the testis but no effect on body weight.  4      only (effects        al. (1982)
    of 10 males, 4    daily for 9 days.             out of 10 animals had isolated atrophic tubules     observed at 2000
    to 6 weeks old)                                 while in the other 6 mice, only 10 to 20% of the    mg/kg b.w. per day)
                                                    tubules showed pronounced atrophy.

    Mouse (dd; 5      3, 9, 27, 50, 100 or 200      Reduction in relative testicular weight and         NOEL = 27 mg/kg      Sajiki
    males and 5       mg/kg b.w. per day in olive   increases in leukocyte counts and serum lactate     b.w. per day         (1975a,b)
    females per       oil by gavage, daily for 30   dehydrogenase activity were observed.               LOEL = 50 mg/kg
    group)            days                          Congestion, oedema and congestive oedema in lung,   b.w. per day
                                                    and loss of spermatogenic cells and spermatogonia
                                                    in testes were observed at doses of >50 mg/kg
                                                    b.w. per day.

    Guinea-pig        2000 mg/kg b.w. per day       There was a decrease in body weight (p<0.05) and    One dose group       Gray et
    (Dunkin-Hartley,  by oral intubation in corn    weight of the testes (p<0.001).  Based on           only (effects        al. (1982)
    groups of 5       oil daily for 7 days.         histological examination of the testes, there was   observed at 2000
    males, 4 to 6                                   severe tubular atrophy with loss of spermatids      mg/kg b.w. per day)
    weeks old)                                      and a reduction in primary spermatocytes and
                                                    spermatogonia.
                                                                                                                                           

    Table 8.  Continued

                                                                                                                                           

    Species                  Protocol                                Results                            Effect Levels        Reference
                                                                                                                                           

    Hamster (males    500 or 1000 mg/kg b.w. per    In the first experiment, the testes, the seminal    In the first         Gray et 
    and females,      day from 20 to 55 days of     vesicles and the epididymis of the high-dose        experiment:          al. (1983)
    strain and        age in one experiment and     group were smaller.  There were no effects on       NOEL = 500 mg/kg     (abstract)
    number            1000 mg/kg b.w. per day from  the female reproductive system.  In the second      b.w. per day 
    unspecified)      20 to 75 days of age in a     experiment, the exposed hamsters had smaller        (males); NOEL =
                      second experiment (nature,    testes; in their offspring, there was decreased     1000 mg/kg b.w. per
                      vehicle and pattern of        viability and growth was retarded.                  day (females)
                      administration unspecified)                                                       In the second
                                                                                                        experiment: one
                                                                                                        dose group only,
                                                                                                        effects observed
                                                                                                        at 1000 mg/kg b.w.
                                                                                                        per day)

    Hamster           2000 mg/kg b.w. per day by    There was no effect on body weight, weight of the   One dose group       Gray et 
    (Syrian DSN,      oral intubation in corn oil   testes or testicular histology.                     only (no effects     al. (1982)
    groups of 7       for 9 days.                                                                       observed at 2000
    males, 4 to 6                                                                                       mg/kg b.w. per day)
    weeks old)
                                                                                                                                           

    Table 8.  Continued

                                                                                                                                           

    Species                  Protocol                                Results                            Effect Levels        Reference
                                                                                                                                           

    Mouse (Swiss      Continous breeding protocol   DBP exposure resulted in a reduction in the         NOEL = 0.3% in the   NTP (1984, 
    CD-1 albino,      with cross-over mating.       numbers of litters per pair and of live pups        diet = 390 mg/kg     1995); Lamb 
    groups of 20      0.03, 0.3 or 1.0% DBP in      per litter, and in the proportion of pups born      b.w. per day.        et al. (1987);
    of each sex)      the diet (equivalent to 39,   alive at the 1.0% amount, but not at lower dose     LOAEL = 1.0% in
                      390, 1300 mg/kg b.w. per      levels.  A cross-over mating trial with the         the diet = 1300
                      day) for a 7-day pre-mating   control and the high dose F0 mice demonstrated      mg/kg b.w. per day
                      period randomly grouped as    that female mice, but not males, were affected
                      mating pairs which were       by DBP, as shown by significant decreases in the
                      exposed during a 98-day       percentage of fertile pairs, the number of live
                      period of cohabitation.       pups per litter, the proportion of pups born
                                                    alive, and live pup weight in the control male
                                                    and exposed female pairing.  In the F0 females,
                                                    absolute and relative liver weights were
                                                    significantly increased and uterine weight was
                                                    significantly decreased at the high dose, which
                                                    suggested that this dose was maternally toxic.
                                                    There were no significant differences in the %
                                                    motile sperm, sperm concentration, or % abnormal
                                                    sperm in the cauda epididymis between male mice
                                                    exposed to 0 or 1.0% DBP in the diet.  No
                                                    treatment-related gross or histopathological
                                                    lesions were noted for the testis, epididymis,
                                                    prostate or seminal vesicles in male mice, or
                                                    for the ovary, oviduct, uterus or vagina in the
                                                    female mice.
                                                                                                                                           
             In a recently reported NTP subchronic study on F-344 rats
    (see NTP rat study 2 in section 7.3), histopathological effects
    in the testes (degeneration of the germinal epithelium) were
    observed at doses of > 720 mg/kg body weight.  The NOEL for
    these effects was 359 mg/kg body weight per day.  In a study with
    combined prenatal perinatal and subchronic exposure (see NTP rat
    study 3 in Section 7.3), similar effects were observed at doses
    of > 571 mg/kg body weight (NOEL = 279 mg/kg body weight per
    day) (NTP, 1995).  In Task 4 of the continous breeding study (see
    Table 8), in Sprague Dawley rats, testicular tubular degeneration
    was observed in the F1 offspring that had been treated during
    the pre- and postnatal period with 322 mg/kg body weight per day,
    and no NOAEL was determined (NTP, 1995; Wine et al., 1997).  The
    F1 offspring showed much more severe damage to the testes and
    secondary sexual organs than did the parent F0 generation at the
    same dose levels.

         The effects on the testes of short-term exposure of rats to
    DBP appear to be at least in part reversible.  Tanino et al.
    (1987) reported that, 2 weeks after discontinuation of the
    administration of 2400 mg/kg body weight per day for 7 days, some
    regeneration of the seminiferous tubules had occurred.  Active
    spermatogenesis was observed in almost all tubules, although
    vacuolation of germinal epithelium and decreased numbers of sperm
    were still evident 3 weeks after exposure had ceased.

         Based on the results of Cater et al. (1977), zinc appears to
    play a role in DBP-induced testicular atrophy in rats.  After 4
    days of exposure of rats to 500, 1000 or 2000 mg DBP/kg body
    weight per day, the weight of the testes was decreased at 
    1000 mg/kg (P < 0.01) and 2000 mg/kg (P < 0.001), and decreased
    further after 6 days (P < 0.05 at 500 mg/kg body weight per day
    and P < 0.001 at the two highest doses) (LOEL = 500 mg/kg body
    weight per day on the basis of decreased testicular weight). 
    Based on histological examination of the testes after 4 days of
    exposure to 2000 mg/kg body weight per day, there was a
    diminution of both spermato-cytes and spermatogonia.  In
    addition, at this dose, urinary zinc excretion was increased by
    34 to 43% over the first 4 days and then returned to normal
    levels.  In the testes, the turnover rate of zinc was enhanced,
    the half-life was decreased from 14 to 5 days and zinc levels
    were significantly reduced; values were 88 - 93% of those of
    controls after 2 days and 64 - 71% after 6 days.  There was no
    change in the turnover rate or zinc content in liver or kidneys. 
    The decrease in the activity of testicular alcohol dehydrogenase
    (a zinc-dependent enzyme) ranged from 20 to 40% of control values
    following 5 days of exposure.  A 3-day pretreatment with
    2000 mg/kg body weight per day caused a 25% decrease (P < 0.001)
    in testicular zinc uptake  in vivo. Concomitant intraperitoneal
    administration of zinc sulfate (50 mg/kg body weight per day)
    resulted in no testicular atrophy (based only on relative testes
    weight rather than on microscopic appearance of the testes).

         Mice and hamsters appear to be somewhat more resistant than
    rats and guinea-pigs to DBP-induced testicular atrophy.  For
    example, testicular effects were observed in rats but not mice in
    NTP subchronic toxicity studies (NTP, 1995).  Following
    administration of 2000 mg DBP/kg body weight per day by gavage in
    corn oil for 7 to 9 days, only isolated tubular atrophy was
    observed in 40% of the mice; no effects on testicular histology
    were observed in hamsters, but oral doses of 2000 mg/kg body
    weight per day administered to rats and guinea-pigs produced
    severe tubular atrophy with loss of spermatids and reductions in
    primary spermatocytes and spermatogonia (Gray et al., 1982).  In
    a study reported in the form of an abstract (Gray et al., 1983a),
    pronounced effects were observed in male rats following
    administration of 500 or 1000 mg/kg body weight per day from 22
    to 55 days of age, while effects in hamsters were observed at the
    high dose only.  Exposed rats and hamsters both had smaller
    testes and seminal vesicles than controls.  The rats also had no
    sperm in the vas deferens and the size of the epididymis of the
    hamster was reduced.  In a separate experiment reported in the
    same abstract (Gray et al., 1983a), half of the rats exposed to
    500 mg/kg body weight per day from 20 to 75 days of age did not
    breed, while hamsters exposed to 1000 mg/kg body weight per day
    had smaller testes but bred, although the viability of their
    offspring was decreased and growth was retarded.  In other
    studies in which ICR male mice ingested a diet containing 2% DBP
    (equivalent to 2400 mg/kg body weight per day), there was a
    decrease in testicular levels of zinc but no testicular atrophy
    was apparent (Oishi & Hiraga, 1980b), whereas testicular atrophy
    and decreased testicular zinc levels were observed in male Wistar
    rats exposed under the same conditions (2% in the diet equivalent
    to 2060 mg/kg body weight per day) (Oishi & Hiraga, 1980a).  In
    an early study, the results of which have not been confirmed at
    such low doses in mice by other investigators, Sajiki (1975a,b)
    reported loss of spermatogenic cells and spermatogonia at doses
    of >50 mg/kg body weight per day administered for 30 days by
    stomach intubation.

         In summary, alteration in testicular enzymes and
    degeneration of testicular germinal cells in  rats have been
    observed at doses of 250, 322 and 571 mg/kg body weight per day,
    respectively.  Effects on a second generation may be more severe
    than on the first generation (Srivastava et al., 1990a, 1990b;
    NTP, 1995).  There are considerable species differences in
    effects on the testes following exposure to DBP, minimal effects
    being observed in mice and none in hamsters at doses as high as
    2000 mg/kg body weight per day (Gray et al., 1982).

         It has been suggested (Foster et al., 1982) that part of the
    difference in sensitivity of the rat and hamster to the
    testicular toxicity of DBP relates to the higher levels of free
    MBP in the rat (with lower levels of conjugate) than in the

    hamster, coupled with the increased testicular ß-glucuronidase
    activity in the rat.  Both of these could lead to higher levels
    of MBP in the rat testis.  No data have been identified on
    metabolism or tissue levels in mice or humans.

         Results of available studies on the effects of MBP (the
    principle metabolite of DBP) on the testes are presented in
    Table 9.  MBP induces testicular damage at doses similar to those
    of DBP.  Clear signs of testicular atrophy were observed after
    oral administration of MBP to rats (Table 9).  There were
    reductions in testicular weight in rats ingesting 400 to 800 mg
    MBP/kg body weight per day for periods of 5 or 6 days (Cater et
    al., 1977; Gray et al., 1980).  Histologically, the majority of
    seminiferous tubules in animals administered 800 mg/kg body
    weight per day for 6 days were atrophied, and there were
    reductions in numbers of spermatocytes and spermatogonia (Foster
    et al., 1981).  Reductions in testicular zinc and relative testes
    weights were observed in rats administered 2% MBP in the diet for
    1 week (Oishi & Hiraga, 1980c).

          In vitro exposure of human spermatozoa to 278 mg DBP/litre
    resulted in a 25% reduction in motility (Fredricsson et al.,
    1993).

    7.5.1.2  Effects on fertility

         Available data concerning the effects of DBP on fertility
    are presented in Table 8.  Corresponding data for MBP are
    presented in Table 9.

         Cummings & Gray (1987) reported that in rats, DBP had no
    effect on early pregnancy during short-term exposure following
    ovulation and continuing throughout the period of implantation
    during pregnancy and pseudopregnancy at doses up to 2000 mg/kg
    body weight per day; neither the number of implantation sites,
    uterine weight, ovarian weight nor serum progesterone
    concentrations were affected (relative to vehicle-exposed
    controls).  In addition, there were no significant effects on the
    decidual cell response.  These data indicated that DBP did not
    affect any maternal parameter of progestational physiology
    including the ability of the uterus to undergo deciduation.

        Table 9.  Reproductive effects of monobutyl phthalate

                                                                                                                                            

    Species                   Protocol                                 Results                            Effect levels      Reference
                                                                                                                                            

    Rat (Wistar,      800 mg/kg b.w. per day         Loss of weight in testes, seminal vesicle and      One dose group       Ikemoto et
    groups of 5       by gavage (vehicle             prostate was evident; an almost complete absence   only; effects        al. (1983)
    males, 5 weeks    unspecified), daily for        of germ cells in seminiferous tubules.             observed at 800
    old)              7 days                                                                            mg/kg b.w. per day

    Rat (Wistar,      800 mg/kg b.w. per day by      Decreases in relative weights of testes, seminal   One dose only;       Ikemote
    groups of 30      gavage in dimethyl sulfoxide   vesicle and prostate were evident.                 effects observed     (1985)
    males, 35 days    for 1 week; animals were       Histopathologically, germ cells had almost         at 800 mg/kg b.w.
    old)              sacrificed one day, 2 weeks    disappeared in seminiferous tubules with           per day
                      and 4 weeks after final        vacuolation and enlargement.  Concentration of
                      dosing.                        testosterone in serum was decreased but FSH and
                                                     LH concentration were not changed.  Four weeks
                                                     after end of dosing, spermatogenesis had
                                                     recovered.

    Rat               800 mg/kg b.w. per day by      Testicular weight was reduced to 57% of control    One dose group       Grey et
    (Sprague-Dawley)  oral intubation                value (p<0.001).  Upon microscopic examination,    only; effects        al. (1982)
                                                     tubular atrophy was observed in all treated rats.  observed at 800
                                                     mg/kg b.w. per day

    Rat               800 mg/kg b.w. per day for     Relative weights of testes and seminal vesicles    One dose group       Foster et
    (Sprague-Dawley,  6 days by oral intubation      were decreased.  Urinary zinc excretion was        only; effects        al. (1981)
    groups of 6                                      increased.                                         observed at 800
    males)                                                                                              mg/kg b.w. per day

    Rat               400 or 800 mg/kg b.w. per      Decrease in testicular weight at both doses after  LOEL = 400 mg/kg     Cater et
    (Sprague-Dawley,  day for 4 or 6 days by         4 days; the weight  decreased further after 6      b.w. per day         al. (1977)
    groups of 6       gavage                         days.
    males)
                                                                                                                                            

    Table 9.  Continued

                                                                                                                                            

    Species                   Protocol                                 Results                            Effect levels      Reference
                                                                                                                                            

    Rat (JCL:Wistar,  2% in the diet for 7 days      Mean body weight was significantly lower than      One dose group       Oishi &
    groups of 10      (2000 mg/kg b.w. per day)      that of control.  Absolute and testicular          only; effects        Hiraga
    males, 5 weeks                                   weights were decreased.  Zinc concentration in     observed at 2000     (1980c)
    old)                                             the testis was decreased.  Testosterone            mg/kg b.w. per day
                                                     concentration in the testis was increased but
                                                     that in serum was not changed.

    Mouse (JCL:ICR,   2% in the diet for 7 days      Body weight gain decreased.  Relative weight of    One dose group       Oishi &
    groups of 10      (2500 mg/kg b.w. per day)      testes was increased.  Concentrations of zinc and  only; effects        Hiraga
    males, 5 weeks                                   testosterone in the testis were decreased.         observed at 2500     (1980d)
    old)                                                                                                mg/kg b.w. per day

    Hamster (DSN,     1600 mg/kg b.w. per day for    Occasional tubular atrophy observed in 2                                Gray et
    groups of 7       9 days by oral intubation      hamsters                                                                al. (1982)
    males)
                                                                                                                                            
             In NTP study 4 (section 7.3), DBP was administered in the
    diet (0, 1000, 5000 or 10 000 mg/kg diet) to Sprague-Dawley rats
    in a continuous breeding protocol, which included cross-over
    mating and offspring assessment phases (NTP, 1995; Wine et al.,
    1997).  Additional details on the protocol and levels at which
    effects occurred are presented in Table 8.  Average equivalent
    dose levels on a body weight basis were 66, 320 and 651 mg/kg
    body weight (NTP, 1991). Mean body weights of exposed dams
    generally decreased with increasing dose and, in the high-dose
    group, were 6 - 13% lower than those of controls at delivery and
    during lactation.  In the F0 generation, the average number of
    live pups per litter (all groups) and mean pup weight at birth
    and during lactation (mid- and high-dose groups) were less than
    in controls.  Cross-over mating trials in the F0 generation
    revealed no effects on the fertility of male or female rats
    receiving the highest dose, although the live pup weight, when
    adjusted for litter size, was significantly less for litters from
    exposed dams. The absolute liver weight of exposed male rats and
    relative liver and kidney weights of exposed male and female F0
    rats of the high-dose group were significantly greater than those
    in controls.  In contrast to the F0 rats, mating, pregnancy and
    fertility indices of F1 rats were lower in the high-dose group
    than in controls (1/20 pregnant versus 19/20 in the controls). 
    In the high-dose group of F1 males, absolute and relative
    epidydimal, right caudal epididymal, right testis, seminal
    vesicle and prostate gland weights were reduced; germinal
    epithelial degeneration of the testes, absence or
    underdevelopment of the epididymides and interstitial cell
    hyperplasia were also observed.  Epididymal sperm count and
    concentration and testicular spermatid head count and
    concentration were also significantly decreased in the high-dose
    group of males. Seminiferous tubule degeneration was observed in
    1/10, 3/10 and 8/10 in the controls, mid- and high-dose groups,
    respectively.  In F1 females, the right ovary weights were
    unchanged. Total and adjusted live pup weights were lower in all
    exposed groups than in the controls.  No clear NOEL was
    established in this study.  In the first generation (F0) the
    reduction in pup weight in the mid-dose group, in the absence of
    any adverse effect on maternal weight, can be regarded as a
    developmental toxicity effect.  There was also a significant
    reduction of live litter numbers at all three dose levels.  The
    effects in the second generation were more severe, with reduced
    pup weight in all groups including the low-dose group, structural
    defects in the mid- and high-dose groups, and severe effects on
    spermatogenesis in the high-dose group that were not seen in the
    parent animals.  These results suggest that the adverse effects
    of DBP are more marked in animals exposed during development and
    maturation than in animals exposed as adults only (Wine et al.,
    1997).

         In NTP rat study 1, DBP was administered in the diet to
    F-344 rat dams during gestation and lactation and to the pups
    postweaning for four additional weeks, at concentrations of 0,
    1250, 2500, 5000, 7500, 10 000 and 20 000 mg/kg diet.  Based on
    decreased weight gains in the dams at >10 000 mg/kg, decreased
    gestation index and increased pup mortality at 20 000 mg/kg,
    decreased body weight of pups at 10 000 mg/kg and mild to marked
    epidydimal hypospermia at >7500 mg/kg, 10 000 mg/kg was
    recommended as the maximum perinatal exposure concentration for
    male and female rats for subsequent studies (NTP, 1995).

         DBP was administered in the diet (0, 300, 3 000 or
    10 000 mg/kg diet) to Swiss (CD-1) mice (NTP, 1995).  Average
    equivalent dose levels on a body weight basis were 0, 39, 390 or
    1300 mg/kg body weight per day (NTP, 1984).  In F0 mice in the
    high-dose group that received DBP during the continuous breeding
    phase, the fertility index, average number of litters per
    breeding pair, live male and female pups, and live pups per
    litter were significantly lower than in the controls.  The ratio
    of live male pups to total live pups in the high-dose group was
    greater than in the controls.  In the cross-over mating trial,
    the fertility index, numbers of live male, female and total pups
    per litter, and total and adjusted live pup weights were
    significantly lower for F0 females (bred with control males) in
    the high-dose group than for the control females bred with
    unexposed males.  The female pup weights in litters from control
    females bred with exposed males were also lower than those of
    control females bred with unexposed males.  Fertility was not
    affected, though the pup weights were lower.  In females in the
    10 000 mg/kg group, the liver weight was greater and the uterine
    weight was less than in control females.  Based on comparison
    with a similar study in rats, mice therefore appear to be less
    sensitive than rats to reproductive effects of DBP, effects only
    being seen at the highest dose level (NTP, 1995).

         In a prenatal/perinatal range-finding study, 20 pregnant
    B6C3F1 mice per group were exposed to 0, 1250, 2500, 5000, 7500,
    10 000 or 20 000 mg DBP/kg diet throughout gestation and
    lactation.  Pups were weaned onto the same diet as their dams and
    exposed for a further 4 weeks (NTP, 1995).  Developmental
    toxicity and pup mortality were suggested at concentrations as
    low as 7500 mg/kg.

         In a study reported only as an abstract, Gray et al.
    (1983a), reported no effects on the female reproductive system in
    an unspecified number of hamsters exposed to 500 or 1000 mg/kg
    body weight per day from 20 to 55 days of age.  In a second
    experiment in the same report, however, half of the breeding
    pairs of rats exposed to 500 mg/kg body weight per day from 20 to
    75 days of age did not breed (NOEL = 250 mg/kg body weight
    per day).  At similar doses (LOAEL = 500 mg/kg body weight per
    day; NOEL = 250 mg/kg body weight per day), breeding in rats was
    adversely affected (Gray et al., 1983a,b).

         Heindel et al. (1989) reported the results of a reproductive
    study for diethylhexyl phthalate in mice.  Though data on the
    other phthalates were not presented in the published report, the
    authors concluded, on the basis of similar studies for these
    compounds, that the relative order of reproductive toxicity for
    the various phthalates was diethylhexyl, dihexyl, dipentyl, di-
     n-butyl and dipropyl; diethyl and dioctyl phthalates were
    considered non-toxic.

    7.5.2  Developmental effects

         The developmental effects of DBP have been examined in rats
    and mice following oral and intraperitoneal administration (the
    latter considered less relevant for assessment of dose-related
    effects), as summarized in Table 10.  DBP generally induced
    fetotoxic effects in the absence of maternal toxicity, and
    teratogenic effects only at high maternally toxic doses.

         Ema et al. (1993) administered DBP by gavage to Wistar rats
    on days 7-15 of gestation at dose levels of 0, 500, 630, 750 and
    1000 mg/kg body weight per day.  No effects in either dams or
    offspring were reported at 500 mg/kg body weight.  At the LOEL of
    630 mg/kg body weight per day, there was a significant increase
    in maternal body weight gain, significantly increased incidence
    of postimplantation loss, and significant decrease in fetal
    weight and increased malformations.  The NOEL was 500 mg/kg body
    weight per day.

         Results of a recent study indicate that susceptibility to
    teratogenesis varies with the developmental stage during the
    period of DBP administration, based on exposure of Wistar rats to
    0, 750, 1000 or 1500 mg/kg on either days 7 to 9, days 10 to 12,
    or days 13 to 15 of gestation (Ema et al., 1994).  When DBP was
    administered  on  days  10  to  12  of gestation, there was no
    evidence of teratogenicity.  Following administration on days 7
    to 9 and 13 to 15, the frequency of malformations increased with
    dose level, and was highest when DBP was administered on days 13
    to 15 (information on maternal toxicity was not reported). 
    Malformations were also observed during the postnatal development
    of the rats of the final litter in the Continuous Breeding
    protocol study (Task 4, NTP, 1995; Wine et al., 1997) (Table 8). 
    Three out of 20 males of the high-dose group that had been
    exposed pre- and postnatally to 650 mg DBP/kg body weight had
    small and malformed prepuces and/or penises and non-palpable
    testes.  Five out of ten rats examined histologically had
    underdeveloped or defective epididymides (Wine et al., 1997).

        Table 10.  Developmental effects of DBP

                                                                                                                                            

    Species                 Protocol                                Results                             Effect levels          Reference
                                                                                                                                            

    Rat (Holtzman,    Pseudopregnant rats          No effect on the decidual cell response,          NOEL = 2000 mg/kg       Cummings &
    groups of 6       received 0, 250, 500,        pregnant uterine weight, number of                b.w. per day            Gray (1987)
    pseudopregnant    1000 or 2000 mg/kg b.w.      implantation sites, ovarian weight, or serum
    females and       per day while pregnant       progesterone concentration during early
    groups of 6 to    rats received 0, 500,        pregnancy or pseudopregnancy.  These data
    8 pregnant        1000 or 2000 mg/kg b.w.      indicated that short-term exposure to DBP had
    females)          per day by gavage in         no direct maternal effects in the rat and
                      sesame oil from day 1        suggested that the viability of preimplantation
                      through day 8 of             embryos was not compromised.
                      pseudopregnancy or
                      pregnancy.

    Rat (females,     250 mg/kg b.w. per day       A large increase in total embryonal death,        One dose group only     Aldyreva
    strain and        by gavage daily (vehicle     owing to high  preimplantation losses, was        (effects observed at    et al. (1975)
    number            unspecified) at various      noted.  Data on maternal toxicity were not        250 mg/kg b.w. per
    unspecified)      stages of gestation or       presented.                                        day)
                      over the first 22 days of
                      gestation.

    Rat (Wistar,      120 or 600 mg/kg b.w. per    No effects on ossification, bone development of   NOEL = 120 mg/kg b.w.   Nikonorow
    groups of 10      day by gavage in corn oil    the base of the skull, paws of the front and      per day (offspring)     et al. (1973)
    or 20 females)    to groups of 20 females      hind extremities or rib fusion in fetuses.        LOEL = 600 mg/kg b.w.
                      prior to or during mating.   Increased number of resorptions and decreased     per day (offspring)
                      120 or 600 mg/kg b.w. per    fetal body weights were observed at the high
                      day by gavage in olive oil   dose.  However, these abnormalities were not
                      daily throughout gestation   observed when DBP was administered prior to and
                      (21 days).                   during mating.  Maternal toxicity was not
                                                   addressed.
                                                                                                                                            

    Table 10.  Continued

                                                                                                                                            

    Species                 Protocol                                Results                             Effect levels          Reference
                                                                                                                                            

    Rat, (Wistar,     Administered DBP by          Significant decrease in maternal body weight      NOEL = 500 mg/kg b.w.   Ema et al.
    groups of 11      gavage in olive oil, days    gain at >630 mg/kg b.w.  Maternal deaths at       per day                 (1993)
    or 12 females)    7-15 of gestation, 0, 500,   1000 mg/kg b.w. per day.  Significant increase
                      630, 750 or 1000 mg/kg       in postimplantation loss at 630 mg/kg b.w. per
                      b.w. per day                 day with complete resorption of implanted
                                                   embryos in surviving dams at 1000 mg/kg b.w.
                                                   per day.  Significant decrease in fetal weight
                                                   at >630 mg/kg b.w.  Increased incidence of
                                                   malformed fetuses (predominantly cleft palate)
                                                   at >630 mg/kg b.w. (significant at 750 mg/kg
                                                   b.w).

    Rat (Wistar,      Pregnant rats were housed                                                                              Ema et al.
    groups of 11      individually and dosed by                                                                              (1994)
    or 12 females)    gastric intubation with DBP
                      (99% pure) in olive oil.
                      All rats killed on day 20.

                      Dosing on days 7 to 9 of     Significant increase in number of resorptions,    LOAEL = 750 mg/kg b.w.
                      pregnancy - 0, 750, 1000 or  dead fetuses per litter; postimplantation         per day (teratogenic
                      1500 mg/kg b.w. per day      loss at >0.75 g/kg b.w. per day  (100%            effects)
                                                   postimplantation loss at 1.5 g/kg b.w. per
                                                   day). Reduction in body weight of male and
                                                   female fetuses at >750 mg/kg b.w. per day.
                                                   Significant increase of fetuses with skeletal
                                                   malformations, fusion or absence of cervical
                                                   vertebral arches or ribs at 750 mg/kg b.w.
                                                   per day.
                                                                                                                                            

    Table 10.  Continued

                                                                                                                                            

    Species                 Protocol                                Results                             Effect levels          Reference
                                                                                                                                            

    Rat (Wistar,      Dosing on days 10 to 12      Significant increase in number of resorptions     LOAEL = 750 mg/kg       Ema et al.
    groups of 11      of pregnancy - 0, 750,       and dead fetuses per litter; and                  b.w. per day (no        (1994)
    or 12 females)    1000 or 1500 mg/kg b.w.      postimplantation loss per litter at >750 mg/kg    teratogenic effects)
                      per day                      b.w. per day (100% postimplantation loss at
                                                   1500 mg/kg b.w. per day).  Reduction in body
                                                   weight of live female fetuses at 750 mg/kg b.w.
                                                   per day and change in sex ratio of live fetuses
                                                   at 1000 mg/kg b.w. per day.

                      Dosing on days 13 to 15      Significant increase in postimplantation loss     LOAEL = 750 mg/kg b.w.
                      of pregnancy - 0, 750,       per litter at >750 mg/kg b.w. per day (100%       per day (teratogenic
                      1000 or 1500 mg/kg b.w.      postimplantation loss at 1500 mg/kg b.w. per      effects)
                      per day                      day) and number of resorptions and dead fetuses
                                                   per litter at 1000 mg/kg b.w. per day.
                                                   Increase in fetuses with malformations, cleft
                                                   palate, skeletal malformations and fusion of
                                                   sternebrae at >750 mg/kg b.w. per day.

    Mouse (ICR-JCL,   0.05, 0.1, 0.2, 0.4 or       There was a significant reduction in body         NOEL = 370 mg/kg b.w.   Shiota et al.
    groups of 7 to    1.0% in the food throughout  weight at day 18 in mothers administered the      per day (offspring)     (1980);
    15 females)       gestation (18 days)          highest dose.  The total number of implants was   LOEL = 660 mg/kg b.w.   Shiota &
                      corresponding to 80, 180,    similar in exposed and control animals but the    per day (offspring)     Nishimura
                      370, 660 and 2100 mg/kg      numbers of resorptions and dead fetuses were      NOEL = 660 mg/kg b.w.   (1982)
                      b.w. per day based on data   much higher in high-dose animals.  There was a    per day (mothers)
                      on food consumption.         dose-dependent decline in fetal body weights
                                                   but this was only significant at the two higher
                                                   doses.  There were no abnormalities except in
                                                   the group exposed to the highest concentration,
                                                   in which there were 2 fetuses (75%) which had
                                                   exencephaly and myeloschisis.  No malformations
                                                   of internal organs were observed in the fetuses
                                                   examined by the microdissection method.
                                                                                                                                            

    Table 10.  Continued

                                                                                                                                            

    Species                 Protocol                                Results                             Effect levels          Reference
                                                                                                                                            

    Mouse (ICR-JCL,   0, 0.005, 0.05, 0.5% DBP     The number of pregnant animals, the incidences    NOEL = 0.05% (100       Hamano et
    groups of 15 to   in the diet                  of spontaneous abortion and maternal deaths,      mg/kg b.w. per day)     al. (1977)
    18 females)                                    and the number of mice with live offspring        (offspring)
                      Based upon food intake       were similar in the exposed groups and            LOAEL = 0.5% (400
                      data, the two highest        controls.  No effects were noted on maternal      mg/kg b.w. per day)
                      doses were calculated to     liver and spleen weights.  A statistically        (offspring)
                      be 100 and 400 mg/kg b.w.    significant increase in kidney weight was         NOEL = 0.05% (100
                      per day                      observed in the high-dose group.  An              mg/kg b.w. per day)
                                                   embryotoxic effect was noted at the highest       (mothers)
                                                   concentration,  resulting in a lower number
                                                   of live offspring.  The incidence of external
                                                   anomalies was also significantly higher in the
                                                   high-dose group.  At the high dose, these
                                                   anomalies were non-closing eyelid (3),
                                                   encephalocoele (6), cleft palate (1), spina
                                                   bifida (1), non-closing eyelid + encephalocoele
                                                   (3).  The rate of ossification for all dosed
                                                   groups appeared to be within normal limits.
                                                   The incidence of skeletal anomalies, especially
                                                   of the sternum, was higher (but not
                                                   statisticaal significant) in the high-dose
                                                   group with respect to the controls.

    Mouse (CD-1,      2500 mg/kg b.w. per day by   5 exposed mice died and there were no viable      One dose group only     Hardin et al.
    group of 50       gavage in corn oil on days   litters.  Maternal toxicity was not addressed.    (effects observed at    (1987)
    females)          6 to 13 of gestation.                                                          2500 mg/kg b.w. per
                                                                                                     day)
                                                                                                                                            

    Table 10.  Continued

                                                                                                                                            

    Species                 Protocol                                Results                             Effect levels          Reference
                                                                                                                                            

    Rat               0, 0.32, 0.64 or 1.06 g/kg   At all doses, the number of resorptions in        LOAEL = 320 mg/kg b.w.  Singh et al.
    (Sprague-Dawley,  b.w. intraperitoneally on    exposed animals was higher than in controls       per day (offspring)     (1972)
    groups of 5       days 5, 10 and 15 of         and there was a corresponding decrease in the
    females)          gestation.                   number of live fetuses.  Fetal weight was
                                                   significantly lower than in controls at all
                                                   doses.  There was a higher incidence (not
                                                   analysed statistically) of skeletal anomalies
                                                   in exposed animals when compared with
                                                   unexposed controls.  These were mainly rib
                                                   abnormalities, absence of tail bones,
                                                   incomplete skull bones and incomplete or
                                                   missing leg bones.  Maternal toxicity was
                                                   not addressed.

    Rat               2 ml/kg b.w. (2080 mg/kg     One rat administered the highest dose died.       LOEL = 2080 mg/kg b.w.  Peters &
    (Sprague-Dawley,  b.w.) or 4 ml/kg b.w.        There were no significant effects on                                      Cook (1973)
    groups of 5       (4170 mg/kg b.w.)            implantation.  The average number of pups
    females)          intraperitoneally on days    weaned per litter was significantly lower in
                      3, 6 and 9 of gestation.     exposed animals compared with controls.  Fetal
                                                   abnormalities were not addressed.
                                                                                                                                            
             In the study reported by Hamano et al. (1977), JCL:ICR mice
    were administered 0.005, 0.05 or 0.5% DBP in food throughout 18
    days of gestation (the two highest doses were calculated on the
    basis of food intake to correspond to 100 and 400 mg/kg body
    weight per day).  There were no significant differences in the
    mortality of maternal mice, the rate of spontaneous abortions or
    the rate of premature births between the control and exposed
    groups.  The highest dose was embryotoxic, resulting in a lower
    number of live offspring.  At this highest dose, an increase in
    kidney weight in mothers was reported, although there were no
    effects on the weights of other organs, body weight gain or
    survival in the mothers.  The frequency of offspring with
    external anomalies was also significantly higher in the high-dose
    group than in controls.  The anomalies consisted mainly of spina
    bifida, exencephaly, cleft palate and open eye.  A small but non-
    significant increase in skeletal anomalies was also seen in the
    high-dose group.  Therefore, the NOEL and LOEL values in this
    study were considered to be 100 and 400 mg/kg body weight per
    day, respectively, on the basis of embryotoxic and teratogenic
    effects.

         In summary, the lowest reported LOAEL for developmental
    effects of DBP was that reported by Hamano et al. (1977), i.e.
    400 mg/kg body weight per day for increases in the number of
    resorptions and dead fetuses in JCL:ICR mice.  The NOEL in this
    study was 100 mg/kg body weight per day.

    7.6  Mutagenicity and related end-points

         The weight of the available evidence indicates that DBP is
    not genotoxic (Table 11).  There are no structural alerts
    indicative of potential reactivity with DNA.  The major metabolic
    pathway involves hydrolysis of one ester linkage to yield MBP and
     n-butyl alcohol, neither of which react with DNA.

         DBP (100 to 10 000 œg/plate) was not mutagenic in any of
    four tester strains of   Salmonella tymphimurium in the presence
    or absence of Arochlor-induced rat or hamster liver S-9 (Zieger
    et al., 1985).  These data are consistent with earlier negative
    Ames test studies (Yagi et al., 1976; 1978; Florin et al., 1980;
    Kozumbo et al., 1982).  In two studies, very weak positive
    responses were reported in the absence of an S-9 metabolic
    activation system (Seed, 1982; Agarwal et al., 1985).  These
    results are questionable because the parent compound clearly does
    not react with DNA.  Similarly, an increase in mutant frequency
    was seen without metabolic activation and at very high cytotoxic
    doses in the L5178Y mouse lymphoma cell assay (NTP, 1995).  It
    should be noted, however, that false positive results are common
    in this assay at cytotoxic concentrations.

        Table 11.  Mutagenicity of DBP from HSE, 1986

                                                                                                                                          

    Species                                    Protocol                                    Results                        Reference
                                                                                                                                          

    Salmonella typhimurium;        Levels of up to 1000 µg/plate in the      Negative.  Full data not reported.         Kozumbo et al.
    TA98, TA100                    presence and absence of S-9.                                                         (1982)

    S. typhimurium;                With and without Aroclor S-9.  Tested     Negative.  Complete data not reported.     Florin et al.
    TA98, TA100, TA1535, TA1537    up to levels that precipitated.                                                      (1980)

    S. typhimurium                 Not specified.                            Negative.  No other information provided.  Yagi et al.
    (strain not reported)                                                                                               (1976, 1978)

    S. typhimurium;                Levels of 100-10 000 µg/plate in          Negative.  Full data were not provided.    Zeiger et al.
    TA98, TA100, TA1535, TA1537    DMSO, with and without S-9 in a                                                      (1982)
                                   preincubation-type assay.

    S. typhimurium;                Levels of 0.013, 0.03 and 0.05 mg/ml      Small dose-related increase in             Seed (1982)
    TA100                          in the 8-azaguanine resistance assay      mutation frequency in the absence of
                                   using a preincubation assay with and      S-9, statistically significant at the
                                   without S-9.                              two highest doses.  Values were
                                                                             increased 1.5  times control levels
                                                                             at the highest dose.

    S. typhimurium;                Test for base-pair substitution or        Spot tests yielded negative responses      Agarwal et al.
    TA98, TA100, TA1535, TA1537,   frameshift-type mutations; spot tests     for all strains.                           (1985)
    TA1538 and TA2637.             with 500 µg per plate.                    "Mildly positive" response in TA100
                                   Dose-response test with 100 to 2000 µg    and TA1535, but not in presence of S9.
                                   per plate, with and without S9
                                   metabolic activation.

    Bacillus subtilis; H17         62.5 µg/ml (limit of solubility)          No inhibition indicative of DNA            Sato et al.
    (Rec+) and M45 (Rec-)                                                    damage; positive controls produced         (1975)
                                                                             clear zones of inhibition.  Test did
                                                                             not appear to have been carried out
                                                                             using S-9.
                                                                                                                                          

    Table 11.  Mutagenicity of DBP from HSE, 1986

                                                                                                                                          

    Species                                    Protocol                                    Results                        Reference
                                                                                                                                          

    Pseudo-diploid Chinese         Concentrations of 0.28, 2.78 and          Negative for SCE and chromosomal           Abe & Sasaki
    hamster cell line (Don)        27.8 mg/ml were tested for ability        aberrations.                               (1977)
                                   to induce chromosome aberrations
                                   and sister chromatid exchange
                                   (SCE).  Ethanol solvent.

    Clonal sub-line of a Chinese   Concentrations up to 0.03 mg/ml           Suspicious or equivocal results for        Ishidate &
    hamster fibroblast cell line.  dissolved in aqueous bovine albumin,      induced chromosomal aberrations.           Odashima (1977)
                                   tested for induction of chromosomal
                                   aberrations.

    Human leucocytes (male         Chromosomal aberrations determined        No increase in frequency of                Tsuchiya &
    derived)                       in 100 human leucocytes following         chromosomal aberrations.                   Hattori (1976)
                                   8-h exposure to 0.03 mg/ml DBP in
                                   whole human blood culture.  This
                                   concentration had been previously
                                   shown to inhibit growth of the
                                   culture cells by 20 to 50%.

    Mouse lymphoma cell line       Cells exposed in suspension to DBP        Increased mutant frequency under           NTP (1995)
    L5178Y                         for 4 h in the presence and absence       non-activation conditions with high
                                   of rat liver S9 metabolic activation.     cytotoxicity.

    Mouse Balb/3T3 cells           In vitro transformation assay.            DBP did not induce the appearance          Litton Bionetics
                                   Concentrations of DBP were 3.4,           of a significant number of                 Inc. (1985)
                                   13.7, 27.5, 55.0 and 82.3 nl/ml.          transformed foci.
                                   Protocol included both positive and
                                   negative controls.
                                                                                                                                          
             DBP did not induce sister-chromatid exchanges (SCE) or
    chromosome aberrations in CHO cells (Abe & Sasaki, 1977) but an
    equivocal result was reported for induction of chromosome
    aberrations in a Chinese hamster fibroblast cell line in the
    absence of metabolic activation (Ishidate & Ohashima, 1977).

         DBP was inactive in the Balb/C-3T3  in vitro 
    transformation assay (Litton Bionetics Inc., 1985).

         In the only identified  in vivo study, analysis of
    peripheral blood samples from male and female mice at the end of
    the 13-week feeding study did not reveal any micronuclei (NTP,
    1995).There was no increase in the numbers of revertants in
     Salmonella typhimurium strains TA98 and TA100 exposed to 50 to
    2000 µg/plate MBP, the  principal  metabolite of DBP, in the
    presence and absence of S9 (Yoshikawa et al., 1983).  Similar
    concentrations were also non-mutagenic in two  Escherichia coli
    strains (WP2 uvr A+ and uvr A-).

    7.7  Carcinogenicity

         A long-term carcinogenicity study for DBP has not been
    conducted, although no tumours were observed in two one-year
    bioassays (Smith, 1953; Nikonorow et al., 1973).

    7.8  Special studies

    7.8.1  Induction of metabolizing enzymes

         Following daily oral administration of 0.01, 0.1 or
    1.0 mmol/kg body weight (2.78, 27.8 or 278 mg/kg body weight) DBP
    by gavage in corn oil to male Sprague Dawley rats (n=20) for
    5 days, there was a 48% increase in the hepatic microsomal
    concentration of cytochrome P-450 at 0.01 mmol/kg body weight and
    28-29% increase in NADPH-cytochrome-reductase activity at 0.01
    and 0.1 mmol/kg body weight (Walseth & Nilsen, 1986).  The
    authors concluded that DBP is a moderate to weak inducer of
    several microsomal enzymes though the reasons for increased
    enzyme activity observed at the lower doses but not at the high
    dose (1.0 mmol/kg body weight per day) were not addressed.

         There were no changes in liver, lung or body weights in male
    Sprague Dawley rats (n=15) exposed to 5.7, 28.5 or 79.8 mg
    DPB/m3 air (0.5, 2.5 or 7.0 mg/kg) 6 h per day for 5 days
    (Walseth & Nilsen, 1984).  There was a small but significant
    increase in the activity of hepatic NADPH-cytochrome-c reductase
    in the group exposed to 5.7 mg/m3.  In contrast to the study of
    Walseth & Nilsen, 1986, effects on the hepatic liver microsomal
    enzymes were not observed.  However, the concentration of
    cytochrome P-450 in the lung decreased in a dose-dependent manner
    to a level of 37% of the control.

         In Sprague Dawley rats (n=5) administered intraperitoneally
    3.8 mmol/kg body weight per day (1058 mg/kg body weight per day)
    for 5 days (Walseth et al., 1982), increases in relative liver
    and lung weights and hepatic microsomal cytochrome P-450 content
    were observed; however there was a decrease in pulmonary
    microsomal cytochrome P-450.  In another study in which albino
    male rats (n=5) were intraperitoneally administered 3.05 ml
    DBP/kg body weight per day (3190 mg/kg body weight per day) and
    killed 18 h or 7 days after the treatment (Seth et al., 1981),
    the activity of aniline hydroxylase was inhibited after 18 h. 
    There was also mild inhibition of aminopyrine- N-demethylase
    activity, but no effects on the activities of glucose-6-
    phosphatase or NADPH-cytochrome-c reductase were observed.  There
    was no effect on the activity of hepatic tyrosine
    aminotransferase activity following the single exposure, but
    there was an increase in the activity of this enzyme following
    daily administration.

    8.  EFFECTS ON HUMANS

    8.1  General population exposure

         Cases of sensitization after exposure to DBP have been
    reported.  A 30-year-old woman developed an axillary dermatitis
    after the use of an anti-perspirant containing DBP (Calnan,
    1975).  Patch testing of the skin was positive with both the
    formulation and with DBP, but not with any of the other
    constituents of the formulation.

         In another reported case, a 32-year-old woman noted pruritis
    and redness in the axillae after changing from her usual
    deodorant spray to a new one (Sneddon, 1972).  Patch testing with
    the original formulation, an alternative deodorant spray, 1%
    paraphenylene-diamine and 3% formalin, was negative, but patch
    testing with the new deodorant and DBP, but not the other
    constituents, was positive.

         In a case reported by Husain (1975), a 44-year-old architect
    noticed a patch of eczema under a plastic watch strap on the left
    wrist and after transferring the watch to the right wrist.  The
    results of patch tests were positive for the plastic watch strip,
    20% colophony, 1% paratertiary butylphenol formaldehyde resin and
    5% DBP.

         Cosmetic products containing 4.5-9% DBP were patch tested on
    50 to 250 individuals per sample and no skin sensitization was
    observed (Brandt, 1985).  No other details were provided.

    8.2  Occupational exposure

    8.2.1  Acute toxicity

         Sandmeyer & Kirwin (1981) reported a case of accidental
    poisoning in which a 23-year-old healthy male worker ingested
    10 g of DBP.  Delayed symptoms were nausea, vomiting, and
    dizziness, followed by headache, pain and irritation of the eyes,
    lacrimation, photophobia and conjunctivitis.  Urinalysis was
    abnormal; the urine was dark yellow in colour with sediment and
    contained numerous erythrocytes and leucocytes with moderate
    numbers of oxalate crystals.  Recovery was gradual within 2 weeks
    and complete after 1 month.

    8.2.2  Epidemiological studies

         Identified data are limited to studies of workers exposed to
    mixtures of phthalates.  These include two cross-section studies
    in which similar neurological symptoms were reported (Milkov et
    al., 1973) (Gilioli et al., 1978) and a cross-sectional
    investigation of reproductive effects (Aldyreva et al., 1975).

         Neurological effects were examined based on clinical
    examinations and self-reported symptoms in a cross-sectional
    study of workers employed in the manufacture of artificial
    leather (Milkov et al., 1973).  The workers were exposed and to
    DBP and also to di(2-ethylhexyl) phthalate, di-iso-octyl-
    phthalate and small amounts of di- n-butyl sebacate, di(2-
    ethylhexyl) sebacate and their respective adipates.  Tricresyl
    phosphate was also present in 10-20% of machines used by various
    workers.  The study group consisted of 147 workers (87 females
    and 60 males), the majority (75%) of whom were less than 40 years
    old.  Pain in the upper and lower extremities, accompanied by
    spasms and numbness, was reported in 57% of those employed for 6
    to 10 years (28 persons) and 82% of those employed for more than
    10 years (65 persons).  These symptoms generally developed after
    6-7 years of employment and the pain became continuous with
    increasing length of employment.  Weakness and pain in the legs
    were usually more noticeable on exercise.  Polyneuritis was noted
    in 47 workers (32 with an autonomic-sensory form and 15 with a
    mixed form) predominantly among those with greater length of
    employment.  Another 22 workers (15%) were reported to have
    "functional disturbance of the nervous system".  Approximately
    50% of the workforce was considered normal by the authors.  The
    study was limited, however, by the lack of comparison of effects
    in the exposed workers with those in an appropriate control
    group.  Moreover, it is difficult to attribute the observed
    effects due to DBP since workers were exposed to a mixture of
    phthalates and other compounds, including tricresyl phosphate
    which is believed to induce polyneuritis.

         A cross sectional study of neurological symptoms based on
    clinical examination was carried out on three groups of male
    workers in Italy who were involved in the production of phthalate
    esters (Gilioli et al., 1978).  The first group of workers was
    exposed to phthalates (23 subjects), while the second and third
    groups were exposed to alcohols (9 subjects) and phthalic
    anhydride (6 subjects), both chemical precursors of phthalate
    esters.  The phthalates involved were di- n-butyl, diisobutyl,
    di(2-ethylhexyl) and dioctyl phthalates.  Mean concentrations of
    phthalates varied from 1 to 5 mg/m3; peak levels were as high as
    61 mg/m3.  Phthalate-exposed workers frequently complained of
    paraesthesia of the upper and lower limbs.  These symptoms became
    continuous with increasing length of employment.  Excessive
    perspiration of the hands and feet and vasomotor irregularity
    indicative of autonomic effects were observed in 3 workers. 
    Neurological examination revealed polyneuropathy in 12 (57%) of
    the workers exposed to phthalates.  In seven workers, bilateral
    painful decreased sensitivity of skin or senses of the hands and
    feet were noted; three had decreased sense of vibrations. 
    Sensory neuropathy was observed in two workers with long-term
    exposure (13 and 18 years) in the alcohol department;

    hyporeflexia was observed in one worker in the phthalic anhydride
    department.  However, the authors suggested that no definite
    conclusions could be drawn from this study because of the small
    number of workers examined.

         Only one study on the reproductive effects of DBP in humans
    has been reported.  In this cross-sectional investigation
    (Aldyreva et al., 1975), workers were reported to have been
    exposed to levels of DBP in excess of the Maximum Allowable
    Concentration (0.5 mg/m3); however, quantitative data were not
    provided. Based on gynaecological examinations of 189 women
    working in processes involving exposure to DBP, approximately 33%
    were considered to be healthy while 33% were reported to have
    "deviations of the uterus".  The health status of the remaining
    34% was not disclosed.  Decreases in the frequency of pregnancy
    and births were reported in women exposed to phthalates, when
    compared with controls; however, quantitative data on the
    prevalence of effects and the composition of the control group
    were not specified.  There were decreases in the frequency of
    miscarriages (no quantitative data was reported), although this
    probably reflected the decreased frequency of pregnancy.  Based
    on colpocytological examination of 19 of the 189 women, 3 women
    had normal biphasic (oestrogen/progesterone) vaginal cycles,
    2 women had biphasic cycles but with insufficient progesterone
    activity and 3 also had biphasic cycles which were hypohormonal. 
    Anovulatory hypoestrogenous cycles in 10 women and an anovulatory
    hyperestrogenous cycles in 1 woman were observed.  In a control
    group, the composition of which was not described, single-phase
    hyperestrogenous cycles and 2-phase hypoprogesterone cycles were
    common (total incidence 21/28).  The authors believed that these
    results indicated a general increase in progesterone levels in
    exposed women though specific data were not provided to support
    this contention.  It is possible that exposure to DBP may
    contribute to the induction of hormonal changes reflected in
    reduced fertility and changes in the vaginal cycle.  However, on
    the basis of the results from this study, it is difficult to draw
    meaningful conclusions owing to inadequate documentation and lack
    of confirmation of these observations.  In addition, quantitative
    data on exposure of the workers (who were also exposed to a
    variety of other unspecified compounds) to DBP were not provided.

    9.  EFFECTS ON OTHER ORGANISMS IN THE LABORATORY AND FIELD

    9.1  Laboratory experiments

         The results of toxicity studies in various organisms are
    presented in Tables 12 (microorganisms and algae), 13 (aquatic
    invertebrates) and 14 (fish).  Those in which effects were
    observed at lowest concentrations are summarized in the following
    sections.

    9.1.1  Microorganisms

         The toxicity of DBP to microorganisms is summarized in
    Table 12.

         Concentrations of DBP up to 300 mg/litre did not inhibit
    methanogenesis in an anaerobic toxicity assay using secondary
    sludge as the source of a heterogeneous anaerobic population
    (O'Connor et al., 1989).

         In water, the 5- and 30-min EC50 value for DBP was
    10.9 mg/litre in the Microtox Test (Tarkpea et al., 1986). 
    Yoshioka et al. (1985) reported a 24-h EC50 (cell proliferation)
    of 2.2 mg DBP/litre for the protozoan,  Tetrahymena pyriformis.

         Based upon the Microtox test, in which reduction in light
    emitted by the luminescent marine bacteria  Photobacterium
     phosphoreum is determined, the 15-min apparent effects
    threshold (AET)a was estimated to be 1.4 mg DBP/kg dry weight
    for Puget Sound sediment (Tetra Tech Inc., 1986).

              

    a    The AET is defined as the concentration above which
         statistically significant adverse effects are always
         expected relative to appropriate reference conditions.  This
         approach involves comparison of data on the composition of
         sediments collected in contaminated areas to measures of
         biological effects associated with these sediments.  The
         site specificity and lack of assessment of cause-effect
         relationships should be borne in mind when interpreting
         apparent effects thresholds.

        Table 12.  Toxicity of DBP to microorganisms and algae

                                                                                                                                            

    Species                         Study Type             End Point                          Valuea                   Reference
                                                                                                                                            

    Microorganisms

    Bacterium
    (Photobacterium)               Microtox          15-min EC50                      1400 µg/kg dw sediment (m)   Tetra Tech Inc. (1986)
                                                     (reduced luminescence)

    Bacterium
    (Photobacterium phosphoreum)   Microtox          5-min EC50                       10 900 µg/litre (n)          Tarkpea et al. (1986)

    Bacterium
    (Photobacterium phosphoreum)   Microtox          15-min EC50                      11 100 µg/litre (n)          Tarkpea et al. (1986)
                                                     5-min & 30-min EC50              10 900 µg/litre (n)
    Protozoan
    (Tetrahymena pyriformis)       Acute, static     24-h EC50                        2200 µg/litre (n)            Yoshioka et al. (1985)
                                                     (cell proliferation)
                                                                                                                                            

    Table 12.  Continued

                                                                                                                                            

    Species                         Study Type             End Point                          Valuea                   Reference
                                                                                                                                            

    Plants

    Green alga
    (Scenedesmus subspicatus)      Acute, static     48-h EC10 (biomass)              1400 µg/litre (n)            Kühn & Pattard (1990)
                                                     48-h EC50 (biomass)              3500 µg/litre (n)
                                                     48-h EC10 (growth rate)          2600 µg/litre (n)
                                                     48-h EC50 (growth rate)          9000 µg/litre (n)

    Green alga
    (Selenastrum capricornutum)    Acute, static     96-h EC50 (growth inhibition)    750 µg/litre (m)             CMA (1984)
                                                     96-h EC50 (TLM) (survival rate)  20-600 µg/litre (n)          Wilson et al. (1978)
                                                     96-h EC50 (growth)               3.4-200 µg/litre (n)

    Green alga
    (Selenastrum capricornutum)    Chronic, static   10-day EC30 (dec. cell           750 µg/litre (m)             Springborn Bionomics
                                                     numbers)                                                      (1984a)

    Green alga
    (Selenastrum capricornutum)    Chronic, static   7-day NOEL (dec. biomass)        2800 µg/litre (n)            Melin & Egnéus (1983)
                                                     7-day LOEL (dec. biomass)        28 000 µg/litre (n)

                                                                                                                                            
    a    n = nominal concentration; m = measured concentration
        9.1.2  Aquatic organisms

    9.1.2.1  Algae

         The toxicity of DBP to algae is summarized in Table 12.

         A 96-h EC50 (decreased cell numbers) of 750 µg DBP/litre
    was reported for the green alga  Selenastrum capricornutum 
    (Springborn Bionomics, 1984a).  Kühn & Pattard (1990) reported
    48-h EC10 and EC50 values for DBP of 1400 µg/litre and
    500 µg/litre, respectively, for  Scenedesmus subspicatus, based
    on biomass. The 96-h EC50 values for the marine dinoflagellate
     Gymnodinium breve were 3.4-200 µg DBP/litre based on growth and
    20-600 µg DBP/litre based on survival (Wilson et al., 1978). 
    However, care must be taken when interpreting these data because
    the two sets of values are not ranges but each are two replicates
    with large variations.

         Yan et al. (1995) report 96-h EC50 values for dimethyl and
    diethyl phthalate, based on inhibition of growth of  Chlorella
     pyrenoidosa; however, an EC50 for DBP could not be generated
    within the range of its water solubility (13 mg/litre).

         At concentrations of about 7000 µg/litre and 2800 µg/litre,
    DBP reduced the growth rates of the monodispersed marine plankton
    (i.e. non-aggregated)  Thalassiosira pseudomona (diatom) and
     Dunaliella  parva (green algae), respectively (Acey et al.,
    1987).

    9.1.2.2  Invertebrates

         Acute and chronic toxicity data for aquatic invertebrates
    are summarized in Table 13.

         The most sensitive aquatic invertebrates, based on acute
    toxicity tests, are the Mysid shrimp,  Mysidopsis bahia, with a
    96-h LC50 of 750 µg/litre (EG&G Bionomics, 1984a), and the midge
     Chironomus  plumosus, with a 48-h EC50 (based on
    immobilization) of 760 µg DBP/litre (Streufert et al., 1980).

         For chronic studies the most sensitive species in a standard
    test is  Daphnia magna with a 21- day NOEC, based on parent
    survival, of 500 µg/litre (measured value) (Kühn et al., 1989). 
    Another sensitive invertebrate species is the scud,  Gammarus
     pulex, with a 10-day LOEL of 500 µg DBP/litre and a NOEC of
    100 µg/litre, based on reduced locomotor activity (Thurén & Woin,
    1991). A 7-day EC50 of 540 µg DBP/litre has been reported for
    the planarian  Dugesia japonica, based on reduced head
    regeneration (Yoshioka et al., 1986).

        Table 13.  Toxicity of DBP to aquatic invertebrates

                                                                                                                                      

    Species                             Study type         End-point             Concentrationsa             Reference
                                                                                                                                      

    Water flea
    (Daphnia magna)                    Acute, static       48-h LC50             5200 µg/litre         McCarthy & Whitmore (1985)

                                       Acute, static       48-h EC50             3400 µg/litreb        CMA (1984)

                                       Acute, static       48-h LC50             3700 µg/litreb        Call et al. (1983)
                                       renewal

                                       Acute, static       24-h EC50             17 000 µg/litre       Kühn et al. (1989)

                                       Acute, static       24-h EC0              8900 µg/litre

    Grass shrimp
    (Palaemonetes pugio)               Acute, static       96-h LC0              1000 µg/litre         Clark et al. (1987)
                                                           (water
                                                           exposure)

                                                           96-h LC50             >1000 µg/litre
                                                           (water
                                                           exposure)

                                                           96-h LC0              10 mg/kg
                                                           (sediment
                                                           exposure)

                                                           96-h LC50             > 10 mg/kg            Clark et al. (1987)
                                                           (sediment
                                                           exposure)
                                                                                                                                      

    Table 13.  Continued

                                                                                                                                      

    Species                             Study type         End-point             Concentrationsa             Reference
                                                                                                                                      

    Grass shrimp
    (Palaemonetes pugio)                                   10-day LC0            10 mg/kg
                                                           (sediment
                                                           exposure)

                                                           10-day LC50           > 10 mg/kg
                                                           (sediment
                                                           exposure)

    Mysid shrimp                       Acute, static       96-h LC50             750 µg/litre          EG & G Bionomics (1984a)
    (Mysidopsis bahia)

    Scud                               Acute, static       24-h LC50             7000 µg/litre         Mayer & Sanders (1973)
    (Gammarus pseudolimnaeus)
                                                           96-h LC50             2100 µg/litre

    Brine shrimp                       Sublethal, static   72-h NOEL             < 10 000 µg/litre     Sugawara (1974a, 1974b)
    (Artemia salina)                                       (egg hatching,
                                                           larvae survival)

                                       Acute, static       24-h LC50             8000 µg/litre         Hudson et al. (1981)

                                       Acute, static       24-h LC50             5600 µg/litre         Hudson & Bagshaw (1978)

    Crayfish                           Acute, static       24-h LC50 +           > 10 000 µg/litre     Mayer & Ellersieck (1986)
    (Orconectes nais)                                      96 h LC50             

    Harpacticoid                       Acute, static       96-h LC50             1700 µg/litre         Lindén et al. (1979)
    (Nitocra spinipes)
                                                                                                                                      

    Table 13.  Continued

                                                                                                                                      

    Species                             Study type         End-point             Concentrationsa             Reference
                                                                                                                                      

    Midge larvae                       Acute, static       48-h LC50             5400 µg/litre         Mayer & Ellersieck (1986)
    (Chironomus plumosus)
                                                           48-h LC50             4000 µg/litre

                                                           48-h EC50             760 µg/litre          Streufert et al. (1980)
                                                           (immobilization)

    Midge                              Acute, static       48-h EC50             5800 µg/litre         EG & G Bionomics (1984b)
    (Paratanytarsus parthenogenica)

    Benthic community composition      Chronic,            14-day LOEL           340 µg/litre          Tagatz et al. (1983)
                                       flow-through        (decrease number
                                                           of amphipods)

    Planarian                          Acute, static       7-day EC50 (head      540 µg/litre          Yoshioka et al. (1986)
    (Dugesia japonica)                                     regeneration)

                                       Acute, static       7-day LC50            840 µg/litre
                                                           (increased
                                                           abnormalities)

    Water flea                         Chronic,            21-day NOEL           960 µg/litreb         CMA (1984)
    (Daphnia magna)                    flow through        (mortality)

                                                           21-day LOEL           2500 µg/litreb
                                                           (mortality)

                                       Chronic, static     16-day LOEL           1800 µg/litre         McCarthy & Whitmore (1985)
                                       renewal             (survival and
                                                           reproduction)
                                                                                                                                      

    Table 13.  Continued

                                                                                                                                      

    Species                             Study type         End-point             Concentrationsa             Reference
                                                                                                                                      

    Water flea                         16-day NOEL         560 µg/litre
    (Daphnia magna)                                        (survaval and
                                                           reproduction)

                                                           21-day LC50           1920 µg/litreb        DeFoe et al. (1990)

                                                           21-day EC50           1640 µg/litreb
                                                           (reproduction)

                                                           21-day EC50           1050 µg/litreb
                                                           (reproduction)

                                       Chronic, static     21-day NOEL           500 µg/litreb         Kühn et al. (1989)
                                       renewal             (parent survival)

                                       Chronic, static     21-day LOEL           2500 µg/litreb        Springborn Bionomics (1984b)
                                       renewal             (survival and
                                                           reproduction)

    Grass shrimp                       Chronic, static     10-day LC0            10 mg/kg              Clark et al. (1987)
    (Palaemonetes pugio)                                   (sediment
                                                           exposure)

                                                           10-day LC50           > 10 mg/kg
                                                           (sediment
                                                           exposure)

                                       Chronic,            28-day LOEL           1000 µg/litre         Laughlin et al. (1978)
                                       semi-static         (survival)

                                                           28-day NOEL           500 µg/litre
                                                           (survival)
                                                                                                                                      

    Table 13.  Continued

                                                                                                                                      

    Species                             Study type         End-point             Concentrationsa             Reference
                                                                                                                                      

    Scud (Gammarus pulex)              Chronic,            10-day LOEL           > 500 µg/litre        Thurén & Woin (1991)
                                       flow-through        (survival)

                                                           10-day LOEL           500 µg/litre
                                                           (decreased
                                                           locomotor activity)

                                                           10-day NOEL           100 µg/litre
                                                           (decreased
                                                           locomotor activity)

    Midge (Chironomus plumosus)        Chronic,            30-day LOEL           > 560 µg/litre        Streufert & Sanders (1977)
                                       flow-through        (larval emergence)
                                                                                                                                      

    a    all concentrations are nominal unless stated otherwise
    b    measured concentration
             In both laboratory and field studies with estuarine benthic
    systems, DBP had statistically significant effects on 8-week
    colonization at 1000 mg/kg sediment, the highest nominal
    concentration tested (Tagatz et al., 1986).  In the laboratory
    study, the total number of species per box was significantly
    decreased by DBP, while in the field study, only the total number
    of individual molluscs was affected.  The actual exposure
    concentrations were lower than the nominal concentrations, as
    only 48% and 19% of the original concentration persisted in the
    laboratory and field systems, respectively, during the last two
    weeks of the study. In an earlier study in which DBP was
    introduced into the water rather than into the sediment, Tagatz
    et al. (1983) reported that colonization was significantly
    reduced at concentrations of 3700 and 3800 µg/litre in
    laboratory- and field-colonized communities, respectively.  At
    340 µg/litre, there was no statistically significant effect on
    the total numbers of species or individuals in the laboratory-
    colonized community, except that the number of  Corophium
     acherusicum (amphipods) was significantly reduced.  At
    450 µg/litre, DBP did not have a statistically significant effect
    on the field-colonized community.

    9.1.2.3  Vertebrates

         Acute and chronic toxicity data for fish are summarized in
    Table 14.

         In acute toxicity tests the yellow perch ( Perca flavescens)
    and the channel catfish ( Ictalurus punctatus) were the most
    sensitive freshwater fish with 96-h LC50 values of 350 and
    460 µg DBP/litre respectively (Mayer & Ellersieck, 1986).  The
    sheepshead minnow,  Cyprinodon variegatus, for which a 96-h
    LC50 of 600 µg/litre has been reported (CMA, 1984), was the most
    sensitive marine fish species identified.

         Yoshioka et al. (1986) reported a 48-h LC50 of 630 µg
    DBP/litre for the red killifish,  Oryzias latipes, while the
    96-h LC50 values were 730 µg DBP/litre in the bluegill
    ( Lepomis  macrochirus) (Mayer & Sanders, 1973) and 850 µg
    DBP/litre in the fathead minnow ( Pimephales promelas) (DeFoe
    et al., 1990).

         The most sensitive chronic study was based on the rainbow
    trout ( Oncorhynchus mykiss) in an early life stage test with a
    99-day no-observed-effect concentration (NOEC) (growth) of
    100 µg/litre, a 99-day LOEC of 190 µg/litre (growth reduced by
    about 27%) and 100% mortality on day 40 at  400 µg/litre (Ward &
    Boeri, 1991). In an early life stage test on fathead minnows
    ( Pimephales promelas) a 20-day NOEC, based on hatching rate
    and larval survival, of 560 µg/litre was reported (McCarthy &
    Whitmore, 1985).

        Table 14.  Toxicity of DBP to fish

                                                                                                                                       

    Species                       Study type               End-Point            Concentration                  References
                                                                                (µg/litre)a
                                                                                                                                       

    Fathead minnow            Acute, static           24-h LC50                    3300                 Mayer & Ellersieck (1986)
    (Pimephales promelas)
                              Acute, static           24-h LC50                    3000                 EG & Bionomics (1983a)

                              Acute, flow-through     24-h LC50                    4800                 Mayer & Ellersieck (1986)

                              Acute, flow-through     24-h LC50                    1600                 EG & G Bionomics (1983b)

                              Acute                   48-h LC50                    1490                 Mayer & Sanders (1973)

                              Acute, static           48-h LC50 + 96 h LC50        3000                 EG & Bionomics (1983a)

                              Acute, flow-through     48-h LC50                    1200                 EG & Bionomics (1983b)

                              Acute, static           96-h LC50                    2020                 McCarthy & Whitmore (1985)

                              Acute, static           96-h LC50                    1300                 Mayer & Ellersieck (1986)

                              Acute, flow-through     96-h LC50                    850b                 DeFoe et al. (1990)
                              commercial phthalate

                              Acute, flow-through     96-h LC50                    1100b                DeFoe et al. (1990)
                              syntherised phthalate

                              Acute, flow-through     96-h LC50                    3950                 Mayer & Ellersieck (1986)
                              Acute, flow-through     96-h LC50                    920                  EG & Bionomics (1983b)

                              Acute, flow-through     96-h LOEL                    1800                 McCarthy & Whitmore (1985)
                                                      (embryo survival)

                                                      96-h NOEL                    1000
                                                      (embryo survival)
                                                                                                                                       

    Table 14.  Continued

                                                                                                                                       

    Species                       Study type          End-Point                    Concentration               References
                                                                                   (µg/litre)a
                                                                                                                                       

    Yellow perch              Acute, flow-through     24-h LC50                    >1240                Mayer & Ellersieck (1986)
    (Perca flavescens)
                                                      96-h LC50                    350

    Bluegill                  Acute, static           24-h LC50                    1230                 Mayer & Sanders (1973)
    (Lepomis macrochirus)
                                                      24-h LC50                    2100                 Buccafusco et al. (1981)

                                                      24-h LC50                    >3000                Mayer & Ellersieck (1986)

                                                      24-h LC50                    1000                 EG & G Bionomics (1983c)

                                                      48-h LC50                    1200

                                                      96-h LC50                    1200                 Buccafusco et al. (1981)

                                                      96-h LC50                    730                  Mayer & Sanders (1973)

                                                      96-h LC50                    2100 at pH 6.5       Mayer & Ellersieck (1986)

                                                      96-h LC50                    1580 at pH 7.5       

                                                      96-h LC50                    2050 at pH 9.0       Mayer & Ellersieck (1986)

                                                      96-h LC50                    850                  EG & G Bionomics (1983c)

                                                      96-h LC50                    1550                 Mayer & Ellersieck (1986)
                                                                                                                                       

    Table 14.  Continued

                                                                                                                                       

    Species                       Study type                 End-Point             Concentration               References
                                                                                   (µg/litre)a
                                                                                                                                       

    Channel catfish           Acute                   24-h LC50                    3720                 Mayer & Sanders (1973)
    (Ictalurus punctatus)
                                                      96-h LC50                    2910                 Mayer & Sanders (1973)

                              Acute, flow-through     96-h LC50                    460                  Mayer & Ellersieck (1986)

    Rainbow trout
    (Oncorhynchus mykiss)     Acute, static           24-h LC50                    > 16 000             Mayer & Ellersieck (1986)

                              Acute, static           24-h LC50                    2800

                              Acute, flow-through     24-h LC50 + 48-h LC50        1600                 EG & G Bionomics (1983d)

                                                      24-h LC50                    4200                 Mayer & Ellersieck (1986)

                              Acute, static           96-h LC50                    2560

                                                      96-h LC50                    1200                 Hrudey et al. (1976)

                                                      96-h LC50                    6470                 Mayer & Sanders (1973)

                              Acute, flow-through     96-h LC50                    1480                 Mayer & Ellersieck (1986)

                              Acute, flow-through     96-h LC50                    1600                 EG & G Bionomics (1983d)

                              Acute, flow-through     24-h LC50 + 96-h LC50       > 1240                Mayer & Ellersieck (1986)
                                                      (yolk-sac fry)

    Red killifish             Acute, static           48-h LC50                    630                  Yoshioka et al. (1986)
    (Orizias latipes)
                                                                                                                                       

    Table 14.  Continued
                                                                                                                                       

    Species                       Study type                 End-Point          Concentration                  References
                                                                                (µg/litre)a
                                                                                                                                       

    Sheephead minnow          Acute, flow-through     96-h LC50                    600b                 CMA (1984)
    (Cyprinodon variegatus)

    Fathead minnow            Chronic, flow-through   20-day EC100                 1800                 McCarthy & Whitmore (1985)
    (Pimephales promelas)                             (embryo mortality)

                                                      20-day LOEL (hatching        1000
                                                      rate and larval survival)

                                                      20-day NOEL (hatching        560
                                                      rate and larval survival)

    Rainbow trout             Chronic, flow-through   99-day NOEL (growth)         100b                 Ward & Boeri (1991)
    (Oncorhynchus mykiss)
                              Chronic, flow-through   99-day LOEL (growth)         190b

                              Chronic, flow-through   40-day LC100                 400b                 Ward & Boeri (1991)

    Cyprinodontiform fish     Chronic, static         147-day 13% reduction        2000                 Davis (1988)
    (Rivulus marmoratus)                              in embryonic viability

                              Chronic, static         147-day 155% increase        1000
                                                      in skeletal abnormalities
                                                      in progeny of exposed fish

                              Post-exposure           63-day NOEL                  1000                 
                                                      (reproduction)

                              Post-exposure           63-day LOEL                  2000
                                                      (reproduction)
                                                                                                                                       

    a    Concentrations are nominal unless stated otherwise
    b    Measured concentrations
        9.1.3  Terrestrial organisms

    9.1.3.1  Plants

         DBP vapour from flexible plastics (e.g., glazing strips)
    used in greenhouses has been implicated in development of damage
    to plants.  The threshold concentration for visible damage in
    summer cabbage,  Brassica oleracea L. cv. Derby Day was between
    0.141 and 0.360 µg DBP/m3, the latter figure determined in a
    4-week laboratory experiment in which growth restriction,
    chlorosis and cotyledon death were observed (Hardwick et al.,
    1984).  However, it is unclear whether other phthalates
    influenced the observed toxicity.

         At higher concentrations in air, DBP caused damage in other
    plant species.  In strong light, leaves of radish seedlings
    ( Raphanus  sativus) faded to pale green due to the
    disappearance of carotenoid and chlorophyll pigments after 6 days
    exposure to 41.3 to 62.3 µg DBP/m3 air and to white after 9 days
    exposure to 56.5 to 90.7 µg DBP/m3 (Virgin, 1988).  Such effects
    were not observed in wheat seedlings ( Triticum aestivum) when
    exposed to DBP vapour alone, but they did develop when the
    seedlings were also treated with DBP-saturated water.

         DBP inhibited photosynthesis in radish plants ( Raphanus
     sativus) exposed to 120 µg DBP/m3 at a rate of 0.003 m3/min
    for 13 days (Millar and Hannay, 1986).  Concentrations of DBP as
    low as 10 µmoles/m3 (approximately 2800 µg/m3) reduced
    uncoupled electron transport in isolated spinach thylakoids by
    about 13%, while 44 µmoles/m3 (approx. 12 250 µg/m3) caused a
    50% reduction.  Basal electron transport rates were reduced by
    50% at 87 µmoles/m3 (approx. 24 200 µg/m3).

          Application of DBP to leaves of white mustard ( Sinapis
     alba) at a rate of 1.5 µg/cm2 caused chlorosis in new leaves
    as they appeared on the third day after treatment (Lokke &
    Bro-Rasmussen, 1981). This effect did not occur with DBP
    application to nipplewort ( Lapsana communis) or to milfoil
    ( Achillea millefolium).  Plants can also be adversely affected
    by exposure to DBP in the soil.  DBP concentrations in soil of
    200 mg/kg or more reduced the germination of soybeans ( Glycine
     max) by > 33% and decreased the growth of corn ( Zea mays)
    and soybeans by 29 to 80% (Overcash et al., 1982).  Plant height
    and shoot weight were significantly reduced by 17 and 25%,
    respectively, when corn seeds were planted in soil containing
    2000 mg DBP/kg and grown for 3 weeks.  Growth was not affected at
    a concentration in soil of 200 mg DBP/kg (Shea et al., 1982).

         A concentration of 1000 mg DBP/litre (added as a methanol
    solution) reduced seed germination by 48% in peas ( Pisum
     sativum) and by 58% in spinach ( Spinacia oleracea) grown in
    tap water, but there was no observable effect on subsequent
    development of the seeds that did germinate (Herring & Bering,
    1988).  It should be noted, however, that this concentration is
    many times higher than the saturation concentration of DBP in
    water (about 10 mg/litre).

    9.1.3.2  Invertebrates

         The LC50 for DBP in the earthworm  Eisenia fetida was
    1360 µg/cm2 in a 2-day contact test in which the chemical was
    applied to filter paper (the toxic units referring to the amount
    of chemical per cm2 of paper). In comparison, the LC50 of 
    dimethylphthalate was 550 µg/m2 and that of 2,4-dinitrophenol
    was 0.6 µg/m2 (Neuhauser et al., 1985, 1986).

         DBP applied to female house flies topically or by injection
    at a concentration of 20 µg/fly (1000 µg/g body weight) was not
    toxic, causing a mortality of less than 16% after 24 h (Al-Badry
    & Knowles, 1980).  Antagonistic interactions were observed when
    flies were treated simultaneously with DBP and various
    organophosphate insecticides, while synergistic interactions were
    observed when flies were pretreated with the phthalate 30 min
    before exposure to the pesticides.  DBP inhibited the metabolism
    of organophosphate pesticides, accounting for the synergistic
    effects.  When the phthalate and insecticides were applied
    simultaneously, the resulting increase in the total lipophilic
    pool by DBP may have resulted in an internal concentration of
    insecticide below the toxicity threshold.

    9.1.3.3  Vertebrates

         Hill et al. (1975) found no deaths among 10-day-old mallard
    ( Anas platyrhynchos) fed up to 5000 mg DBP/kg for 5 days,
    followed by 3 days on a normal diet.

         In a study in which ring doves ( Streptophelia risoria) 
    were fed a diet containing 10 mg DBP/kg (1.1 mg DBP/kg body
    weight per day) for a period of 3 weeks prior to mating through
    completion of a clutch of two eggs, there was a 23% increase in
    water permeability and a 10% decrease in egg- shell thickness
    (Peakall, 1974). A 15% decrease in shell thickness is considered
    significant for reproductive effects.  Rapid recovery occurred
    upon cessation of exposure.

         Korhonen et al. (1983) studied the embryotoxicity of DBP to
    white leghorn chicken eggs.  On the third day of incubation, DBP
    was injected on the inner shell membrane at doses of 13 and
    26 µmol per egg (3.62 and 7.24 mg/egg, respectively).  At 26 µmol
    per egg, 30 eggs were tested and there were 6 early deaths (2
    days after injection) and 4 non-malformed and 1 malformed late
    deaths (between 3 and 11 days after injection).  An approximate
    ED50 of 33 µmol (9.19 mg) per egg was calculated for DBP.

    10.  EVALUATION OF HUMAN HEALTH RISKS AND EFFECTS ON THE ENVIRONMENT

    10.1  Evaluation of human health risks

    10.1.1  Exposure

         Based on the limited data available, the principal media of
    exposure to DBP for the general population, listed in order of
    their relative importance based upon estimated intake, are as
    follows:  food, indoor air and drinking-water.  Estimated intakes
    from food and indoor air are 7 µg/kg body weight per day and
    0.42 µg/kg body weight per day, respectively.  Estimated intakes
    from drinking-water and ambient air are considerably less,
    < 0.02 µg/kg body weight per day and 0.26-0.36 ng/kg body weight
    per day, respectively. Based on these intakes, it is estimated
    that the total average daily intake from air, drinking-water and
    food is 7.4 µg/kg body weight per day.  It should be noted,
    however, that intake of DBP in the diet can vary considerably,
    depending upon the nature and extent of packaged food consumed
    and the nature of use of food wrapping in food preparation.  In
    the United Kingdom, the Ministry of Agriculture, Fisheries and
    Food has estimated that the maximum likely human intake of DBP
    from food sources is approximately 2 mg per person per day
    (approximately 31 œg/kg body weight per day, assuming a mean body
    weight of 64 kg).  There is also potential for exposure to DBP in
    cosmetics, although available data are inadequate to quantify
    intake from this source.

         The most recent provisional data from the NIOSH National
    Occupational Exposure Survey indicate that in the USA over
    500 000 workers, including 200 000 women, are potentially exposed
    to DBP. At a limited number of worksites in the USA,
    concentrations were generally less than the limit of detection
    (i.e., 0.01 to 0.02 mg/m3).

    10.1.2  Health effects

         The acute toxicity of DBP in mice and rats is low.

         In a case of accidental poisoning of a worker who ingested
    approximately 10 grams of DBP, recovery was gradual within 2
    weeks and complete after 1 month.

         Based on limited available data in animal species, DBP
    appears to have little potential to irritate skin or eyes,
    although in humans a few cases of sensitization after exposure
    have been reported.

         Available data on the effects of DBP in humans are limited
    to those of workers exposed to mixtures of phthalates and are
    inadequate to serve as a basis for assessment of effects of DBP. 
    The remainder of this evaluation is, therefore, based on studies
    in animals.

         The profile of effects following exposure to DBP is similar
    to that of other phthalate esters, which, in susceptible species,
    induce hepatomegaly and increase numbers of hepatic peroxisomes,
    are fetotoxic, have teratogenic potential and produce testicular
    damage.

         Adequate carcinogenesis bioassays for DBP have not been
    conducted.  The weight of  available evidence indicates that DBP
    is not genotoxic.

         As a class, chemicals which cause peroxisome proliferation
    are often hepatocarcinogenic via a non-genotoxic mode of action. 
    Although the mechanism of action remains unknown, tumour
    formation is preceded by peroxisomal proliferation and
    hepatomegaly.  As a chemical causing peroxisomal proliferation,
    it is possible that DBP might be a rodent liver carcinogen,
    although it is much weaker in inducing hepatomegaly and
    peroxisome proliferation than DEHP.  To the degree that
    hepatomegaly and peroxisomal proliferation correlate with
    carcinogenic potency, DBP would be anticipated to be a less
    potent carcinogen than DEHP and would probably exhibit no
    activity as measured by current cancer bioassay methodologies. 
    Thus, it is unlikely that DBP presents any significantly
    increased risk of cancer at concentrations generally present in
    the environment.

         Effects of DBP observed at lowest doses in repeated dose
    toxicity studies are those on the liver and testes of the rat and
    include hepatomegaly and peroxisome proliferation.  In one study,
    hepatic necrotic changes were also reported.  Effects on the
    testes include decreases in the activities of testicular enzymes
    and, at higher doses, degeneration of the germinal epithelium and
    reductions in testicular zinc levels.  DBP also induces adverse
    effects on fertility, is fetotoxic and induces teratogenic
    effects at high concentrations that are toxic to the dams. 
    Toxicity to the testes is more marked when exposure to DBP occurs
    during development and maturation than when adults only are
    exposed.  Lowest reported effect levels in adequate studies for
    these various effects and their associated no-observed-(adverse)-
    effect levels (NOEL/NOAEL) are summarized in Table 15.

        Table 15.  Effect levels of DBP

                                                                                             

                                           No-          Lowest-
                                        observed-      observed-
                                        (adverse)-     (adverse)-
                                       effect levela     effect
                                                         levelb

    End-point           Species          (mg/kg b.w. per day)          Reference
                                                                                            

    Liver:
      Organ weight      rat,           -               120             Nikonorow
      (relative)        Wistar                                         et al. (1973)

      Peroxisomal       rat,           176             356             NTP (1995),
      proliferation     F-344                                          Study No. 2
      and hepatomegaly

                        rat,           138             279             NTP (1995),
                        F-344                                          Study No. 3

      Necrosis (not     rat,                           250             Murakami et
      confirmed)        Wistar                                         al. (1986a)

    Testis:

      Enzymes           rat,                           250             Srivastava
                        Wistar                                         et al.
                                                                       (1990a,b)

      Histopathological rat,           359             720             NTP (1995), 
      lesions           F-344                                          Study No. 2

                        rat,           279             571             NTP (1995), 
                        F-344                                          Study No. 3
                                                                                             

    Table 15.  contd.

                                                                                             

                                           No-          Lowest-
                                        observed-      observed-
                                        (adverse)-     (adverse)-
                                       effect levela     effect
                                                         levelb

    End-point           Species          (mg/kg b.w. per day)          Reference
                                                                                            

    Reproduction/fertility/            rat,            NI c          66NTP (1995;
    developmental       Sprague-                                       Wine et al., 
                        Dawley                                         1997), 
                                                                       Study No. 4

    Developmental       mouse,         100             400             Hamano et al. 
                        JCL:ICR                                        (1977)
                                                                                             

    a    Each value in this column is either a NOEL or NOAEL
    b    Each value in this column is either a LOEL or LOAEL
    c    NI = not identified
    
    10.1.3  Guidance values

         The following guidance is provided as a potential basis for
    derivation of limits of exposure by relevant authorities.  Since
    ingestion is by far the principal route of exposure to DBP and
    since the toxicological data for other routes of administration
    are insufficient for evaluation, only the  oral route is
    addressed here.  However, the ultimate objective should be
    reduction of total exposure from all sources to less than the
    tolerable daily intake presented below.

         The Task Group considered that the testicular and
    reproductive/developmental effects are the most relevant for
    derivation of guidance values for protection of human health. 
    Increases in liver weight, hepatomegaly and peroxisome
    proliferation were regarded by the Task Group as being
    functional, relating most likely to the metabolism of the
    material, rather than pathological.  Moreover, although hepatic
    necrosis was observed in one strain of rats at 250 mg/kg body
    weight per day, it was not observed in two other strains at much
    higher doses.

         The NOAEL/NOEL values for the end-points considered to be
    most appropriate for derivation of guidance values (i.e.
    developmental and reproductive toxicity) have not been identified
    in the Continuous Breeding study (NTP study 4); the lowest dose
    studied (66 mg/kg body weight per day) is a LOAEL (NTP, 1995;
    Wine et al., 1997).  On the basis of these data, an acceptable
    daily intake (ADI) is derived as follows:


    ADI  =     66 mg/kg body weight per day
                                           

                     1000

         =    0.066 mg/kg body weight per day

         =    66 œg/kg body weight per day

    where:

    *    66 mg/kg body weight per day is the approximate LOAEL for
         developmental and reproductive effects in rats observed in
         the most sensitive studies conducted to date

    *    1000 is the uncertainty factor (×10 for interspecies
         variation, ×10 for interindividual variation, ×10 for lack
         of data on a NOAEL.  A factor of 10 for lack of a NOAEL was
         considered adequate since the effects observed at the lowest
         dose levels were moderate and probably reversible.  The
         severe, possibly irreversible, teratogenic, testicular and
         epididymal effects were only observed at the highest dose
         level tested, which also produced other signs of toxicity. 
         Because DBP is rapidly metabolized and eliminated, with no
         evidence of accumulation in tissues, no additional factor
         was incorporated for lack of data on chronic effects.

    10.2  Evaluation of effects in the environment

    10.2.1  Exposure

         DBP exists widely in the environment, being released during
    production, processing, usage and disposal.  However, it is
    relatively non-persistent in air and surface water.  The most
    important process leading to the elimination of DBP is biological
    breakdown, aerobic degradation being rapid and complete. It would
    be expected to be more persistent in anaerobic sediments.  It is
    moderately adsorbed to soil.

         DBP would be expected to bioaccumulate, based on a log Kow
    of 4.3 to 4.7.  However, it tends to be readily metabolized
    leading to bioconcentration factors lower than predicted.
    Biomagnification in terrestrial animals is unlikely.

         Mean concentrations of DBP in surface water tend to be less
    than 1 µg/litre.  However,  levels in polluted rivers are much
    higher, with values of 12 to 34 µg/litre. Levels in sediment are
    generally less than 1 mg/kg dry weight although in polluted areas
    concentrations of up to 10 mg/kg have been measured.  DBP
    concentrations in sewage sludge range from 0.2 to 200 mg/kg dry
    weight.

    10.2.2  Effects

         A comparison of the results of acute and long-term tests on
    aquatic organisms shows that there is no increase in toxic
    effects with increasing duration of exposure.

         In acute toxicity tests the sensitivity of the different
    trophic levels is similar.  The 48-h and 96-h LC50 and EC50
    values for the most sensitive species are in the range of 350 to
    760 µg/litre for freshwater organisms and 600 to 750 µg/litre for
    marine organisms.

         The most sensitive chronic study was based on the rainbow
    trout; the 99-day no-observed-effect concentration (NOEC) based
    on growth was 100 µg/litre and the lowest-observed-effect
    concentration (LOEC) 190 µg/litre.

           The acute toxicity of DBP to birds is low.

    10.2.3  Risk evaluation

         The lowest reported chronic effect level for dissolved DBP
    in aquatic organisms was 190 µg/litre (99-day LOEC for growth)
    and the lowest NOEC was 100 µg/litre in the same test.  These
    values are at least factors of 190 and 100, respectively, greater
    than the mean surface water concentration of DBP.  Therefore, the
    risk to aquatic organisms from mean DBP concentrations in surface
    water is low.  However, in highly polluted rivers where surface
    water concentrations have been found to be up to 34 µg/litre, the
    ratio between the concentration and the NOEC is only 3.

         There is inadequate data to assess the risk of DBP to
    sediment-dwelling organisms.

         The most likely route of exposure for higher organisms,
    e.g., birds and mammals, is through food intake, in particular
    fish.

         The only acute toxicity test on birds was carried out on
    10-day-old mallards, where a 5-day LC50 of > 5000 mg/kg diet
    was found.  Based on food consumption and body weight, an LD50
    for the mallard of > 2043.5 mg/kg body weight can be calculated. 
    Using this data an estimated LC50 for a fish-eating bird (e.g.,
    kingfisher), based on body weight and food consumption, can be
    calculated.

    LC50 (mg/kg dry weight of diet)

              test species LD50 (mg/kg) œ body wt (kg)
    =                                                 

                              food consumption (kg)

         The estimated LC50 for the fish-eating bird is >
    9350 mg/kg diet.  The highest water concentration (34 µg/litre)
    multiplied by the highest bioconcentration factor (590) gives a
    residue level in fish of 20 mg/kg.  Comparing this value to the
    estimated LC50 value gives a Toxicity Exposure Ratio (TER) of
    > 470.  A TER of less than 1 would give cause for concern; but a
    value of > 470 indicates that the risk to fish-eating birds from
    DBP is very low.

         For mammals, the mink, a terrestrial mammal with a diet
    consisting predominantly of aquatic prey, can be used.  The
    estimated intake for a "worst case" scenario is 3.1 mg/kg body
    weight per day. This is based on an ingestion rate of 155 g per
    day and assumes a diet of 75% fish, a maximum measured
    bioconcentration factor of 590 for the fathead minnow, and a
    maximum concentration of DBP in water of 34 µg/litre.  The

    estimated intake is  considerably less than the no-observed-
    adverse-effect levels in toxicity studies in laboratory mammals
    (i.e. 250 mg/kg body weight per day).

    11.  CONCLUSIONS AND RECOMMENDATIONS FOR PROTECTION OF HUMAN HEALTH
         AND THE ENVIRONMENT

         In laboratory animals, the critical toxic effects of DBP
    were those on development and reproduction at concentrations well
    above those to which people are normally exposed in the general
    environment.  DBP is readily broken down in the environment and
    in the body and shows no tendency to accumulate or to persist in
    any specific tissues or organs.  It is unlikely that there is any
    risk to human health at present levels of exposure in the general
    environment.

         There is inadequate information to assess exposure from use
    in cosmetics.

         The risk to aquatic organisms associated with the present
    mean concentrations of DBP in surface waters is low.  However, in
    highly polluted rivers the safety margin is much smaller.  There
    is inadequate data to assess the risk of DBP to sediment-dwelling
    organisms.  At current levels of exposure, it can be concluded
    that the risk to fish-eating birds and mammals is low.

         The current measures being taken to limit the release of DBP
    into the environment and to control its use in food-packaging
    materials should be maintained.

    12.  FURTHER RESEARCH

         The most sensitive end-points used in determining the
    guidance value were effects on reproduction, both fertility and
    development.  No NOAEL was identified for these effects and the
    results suggest that the adverse effects of DBP are more marked
    in animals exposed during development and maturation than in
    animals exposed as adults only.  Data on effects of exposure
    during the developmental period are very limited and further work
    to identify a NOEL for such exposure is urgently needed.

         The number and quality of studies describing the profile of
    toxicity of DBP and its behaviour in the environment are
    sufficient to make reasonable assessments of potential health
    effects, environmental fate and to set a guidance value for
    limiting human exposure to preclude adverse health effects. 
    Therefore, additional research in these areas is of low priority,
    relative to that for other substances.

         However, one area of concern is that the potential for
    exposure in cosmetics is largely unknown.  It is recommended,
    therefore, that additional data on the use and levels of DBP in
    cosmetics be acquired.  If there is a potential for considerable
    additional exposure from this source, it is recommended that
    controlled studies be conducted to examine the rate of skin
    absorption, dosimetry, metabolism and excretion of DBP in humans.

         In view of the large species differences in some toxic
    effects of DBP on laboratory animals, additional studies on
    kinetics and metabolism in humans or human cells are desirable.

         More research is needed on the effects of DBP on sediment-
    dwelling organisms.

    13.  PREVIOUS EVALUATIONS BY INTERNATIONAL BODIES

         DBP may be used for the manufacture of regenerated cellulose
    film which is intended to or does come into contact with food-
    stuffs.  It may be used as a plasticizer in total amount of
    12.5 mg/dm2 on the side of the film in contact with foodstuffs
    (EEC, 1987).

    REFERENCES

    Abe S & Sasaki M (1977) Chromosome aberrations and sister
    chromatid exchanges in Chinese hamster cells exposed to various
    chemicals. J Natl Cancer Inst, 58(6): 1635-1641.

    Aceves M & Grimalt JO (1993) Large and small particle size
    screening of organic cmpounds in urban air. Atmos Environ,
    B27(2): 251-263.

    Acey R, Healy P, Unger TF, Ford CE, & Hudson RA (1987) Growth and
    aggregation behavior of representative phytoplankton as affected
    by the environmental contaminant di-n-butyl phthalate. Bull
    Environ Contam Toxicol, 39: 1-6.

    Adams WJ, Kimerle RA, & Barnett JW (1992) Sediment quality and
    aquatic life assessment. Environ Sci Technol, 26: 1864-1875.

    Agarwal DK, Lawrence WH, Nunez LJ, & Autian J (1985) Mutagenicity
    evaluation of phthalic acid esters and metabolites in
     Salmonella  typhimurium cultures. J Toxicol Environ Health,
    16(1): 61-69.

    Aitio A & Parkki M (1978) Effect of phthalate esters on drug
    metabolising enzyme activities in rat liver. Arch Int
    Pharmacodyn, 235: 187-195.

    Al-Badry MS & Knowles CO (1980) Phthalate-organophosphate
    interactions: toxicity, penetration, and metabolism studies with
    house flies. Arch Environ Contam Toxicol, 9(2): 147-161.

    Al-Omran LA & Preston MR (1987) The interactions of phthalate
    esters with suspended particulate material in fresh and marine
    waters. Environ Pollut, 46: 117-186.

    Albro PW & & Moore B (1974) Identification of the metabolites of
    simple phthalate diesters in rat urine. J Chromatogr, 94: 
    209-218.

    Aldyreva MV, Klimova TS, Izyumova AS, & Timofievskaya LA (1975)
    [Effect of plasticisers on reproductive function.] Gig Tr Prof
    Zabol, 12: 25-29 (in Russian).

    Antonyuk OK & Aldyreva MV (1973) [Substantiation of the maximum
    allowable concentration of dibutyl phthalate in the air of
    industrial premises.] Gig Trud Prof Zabol, 12: 26-30 (in
    Russian).

    Ariyoshi T, Koga Y, Yoshitake S, Takada K, & Tenda N (1976)
    Metabolism of dibutyl phthalate and the effects of its
    metabolites on animals. Kyushu Yakugokkai Kaiho, 30: 17-22.

    Atlas E & Giam CS (1981) Global transport of organic pollutants:
    ambient concentrations in the remote marine atmosphere. Science,
    211: 163-165.

    ATSDR (1990) Toxicological profile for di-n-butylphthalate.
    Atlanta, Georgia, Agency for Toxic Substances and Disease
    Registry, 109 pp.

    Atwater JW, Jasper SE, Parkinson PD, & Mavinic DS (1990) Organic
    contaminants in Canadian coal wastewaters and associated
    sediments. Water Pollut Res J Can, 25: 187-200.

    Barber ED, Astill BD, Moran EJ, Schneider BF, Gray TJB, Lake BG,
    & Evans JG (1987) Peroxisome induction studies on seven phthalate
    esters. Toxicol Ind Health, 3(2): 7-24.

    Barrick R, Becker S, Brown L, Beller H, & Pastrook R (1988)
    Sediment quality values refinement: 1988 update and evaluation of
    Puget Sound AET. Bellevue, Washington, PTI Environmental
    Services.

    Belisle AA, Reichel WL, & Spann JW (1975) Analysis of tissues of
    mallard ducks fed two phthalate esters. Bull Environ Contam
    Toxicol, 13(2): 129-132.

    Bell FP, Patt CS, Brundage B, Gillies PJ, & Phillips WA (1978)
    Studies on lipid biosynthesis and cholesterol content of liver
    and serum lipoproteins in rats fed various phthalate esters.
    Lipids, 13: 66-74.

    Benckiser G & Ottow JCG (1982) Metabolism of the plasticizer
    di-normal-butylphthalate by  Pseudomonas pseudoalcaligenes under
    anaerobic conditions, with nitrate as the only electron-acceptor.
    Appl Environ Microbiol, 44(3): 576-578.

    BIBRA (1986) A 21 day feeding study of di-n-butyl phthalate to
    rats: Effects on the liver  and liver lipids. Carshalton, Surrey,
    The British Industrial Biological Research Association (Report to
    Chemical Manufacturers Association, Washington, DC) (CMA
    reference PE 28.0-BT-BIB).

    Bornmann G & Loeser A (1956) The reaction of the body to the
    action of various plasticizers. Z Lebensm.unters Forsch, 103:
    413-434.

    Bove JL, Dalven P, & Kukreja VP (1978) Airborne di-butyl and di-
    (2-ethylhexyl)-phthalate at three New York City air sampling
    stations. Int J Environ Anal Chem, 5: 189-194.

    Brandt K (1985) Final report on the safety assessment of dibutyl
    phthalate, and diethyl phthalate. J Am Coll Toxicol, 4(3):
    267-303.

    BUA (1987) Dibutyl phthalate. Frankfurt am Main, German Chemical
    Society, Advisory Committee on Existing Chemicals of
    Environmental Relevance, 71 pp (BUA Report No. 22).

    Buccafusco RJ, Ellis SJ, & LeBlanc GA (1981) Acute toxicity of
    priority pollutants to bluegill ( Lepomis  macrochirus). Bull
    Environ Contam Toxicol, 26: 446-452.

    Burns BG, Musial CJ, & Uthe JF (1981) Novel cleanup method for
    quantitative gas chromatographic determination of trace amounts
    of di-2-ethylhexyl phthalate in fish lipid. J Assoc Off Anal Chem,
    64(2): 282-286.

    California Environmental Protection Agency (1992) PTEAM: 
    Monitoring of phthalates and PAHS in indoor and outdoor air
    samples in Riverside, California. Final Report, Volume II
    (Contract No. A933-144). Research Triangle Park, North Carolina,
    Research Triangle Institute (Submitted to California Air
    Resources Board).

    Call DJ, Brooke LT, Ahmad N, & Richter JE (1983) Toxicity and
    metabolism studies with EPA Priority Pollutants and related
    chemicals in freshwater organisms.  Duluth, Minnesota, US
    Environmental Protection Agency, Environmental Research
    Laboratory, Office of Research and Development, 120 pp (EPA
    600/3-83-095).

    Callahan MA, Slimak MW, Gabel NW, May IP, Fowler CF, Freed JR,
    Jennings P, Durfee RL, Whitmore FC, Maestri B, Mabey WR, Holt BR,
    & Gould C (1979) Water-related environmental fate of 129 priority
    pollutants.  Volume II: Halogenated aliphatic hydrocarbons,
    halogenated ethers, monocyclic aromatics, phthalate esters,
    polycyclic aromatic hydrocarbons, nitrosamines, miscellaneous
    compounds. Washington, DC, US Environmental Protection Agency
    (EPA 440/4-79-029b).

    Calley D, Autian J, & Guess WL (1966) Toxicology of a series of
    phthalate esters. J Pharm Sci, 55: 158-162.

    Calnan CD (1975) Dibutyl phthalate. Contact Dermatitis, 1: 388.

    Camford Information Services Inc. (1992) CPI product profiles -
    Di- n-butyl phthalate. Dons Mills, Ontario, Camford Information
    Services Inc., 3pp.

    Castle L, Mercer AJ, Startin JR, & Gilbert J (1988) Migration
    from plasticized films into foods. 3. Migration of phthalate,
    sebacate, citrate and phosphate esters from films used for retail
    food packaging. Food Addit Contam, 5(1):-9-20.

    Castle L, Mayo A, & Gilbert J (1989) Migration of plasticizers
    from printing inks into foods. Food Addit Contam, 6(4): 
    437-443.

    Cater BR, Cook MW, Gangolli SD, & Grasso P (1977) Studies on
    dibutyl phthalate-induced testicular atrophy in the rat: Effect
    on zinc metabolism. Toxicol Appl Pharmacol, 41(3): 609-618.

    Cautreels W & van Cauwenberghe K (1978) Experiments on the
    distribution of organic pollutants between airborne particulate
    matter and the corresponding gas phase. Atmos Environ, 12:
    1133-1141.

    Cautreels W, van Cauwenberghe K, & Guzman LA (1977) Comparison
    between the organic fraction of suspended matter at a background
    and an urban station. Sci Total Environ, 8(1): 79-88.

    CCREM (Canadian Council of Resource and Environment Ministers)
    (1987) Canadian water quality guidelines. Winnipeg, Manitoba,
    Task Force on Water Quality Guidelines of the Canadian Council of
    Resources and Environment Ministers (now part of the Canadian
    Council of Environment Ministers), pp 3/37, 6/182-6/187.

    Chambon P, Riotte M, Daudon M, Chambon-Mougenot R and Bringuier R
    (1971) Etude du métabolisme des phthalates de dibutyle et de
    diéthyle chez le rat. C R Acad Sci Paris, 273: 2165-2168.

    Chan HS (1975) A study of the transfer processes of phthalate
    esters to the marine environment. Houston, Texas, Graduate
    College of Texas A&M University (Ph.D. Dissertation).

    Chapman P (1989) Current approaches to developing sediment
    quality criteria. Environ Toxicol Chem, 8: 589-599.

    Chauret C, Mayfield CI, & Inniss WE (1995) Biotransformation of
    di- n-butyl phthalate by a psychotrophic  Pseudomonas
    fluorescens (BGW) isolated from subsurface environment. Can J
    Microbiol, 41: 54-63.

    Ching NP, Jham GN, & Subbarayan C (1981) Gas chromatographic-mass
    spectrometric detection of circulating plasticizers in surgical
    patients. J Chromatogr, 222: 171-177.

    City of Toronto (1990) The quality of drinking-water in Toronto:
    Summary report. Toronto, City of Toronto, Department of Health.

    Clark JR, Patrick JM Jr, Moore JC, & Lores EM (1987) Waterborne
    and sediment-source toxicities of six organic chemicals to grass
    shrimp ( Palaemonetes pugio) and amphioxus ( Branchiostoma
    caribaeum). Arch Environ Contam Toxicol, 16(4): 401-407.

    CMA (1984) Generation of environmental fate and effects data base
    on 14 phthalate esters. Summary Report - Environmental studies:
    Phase I. Washington, DC, Chemicals Manufacturers Association,
    Phthalate Esters Program Panel.

    Cocchieri RA (1986) Occurrence of phthalate esters in Italian
    packaged foods. J Food Prot, 49(4): 265-266.

    Cummings AM & Gray LE Jr (1987) Dibutyl phthalate: maternal
    effects versus fetotoxicity. Toxicol Lett, 39(1): 43-50.

    Davis WP (1988) Reproductive and developmental responses in the
    self-fertilizing fish,  Rivulus marmoratus, induced by the
    plasticizer, di- n-butylphthalate. Environ Biol Fishes, 21(2):
    81-90.

    Davis BJ, Maronpot RR, & Heindel JJ (1994a) Di-(2-
    ethylhexyl)phthlate suppresses estradiol and ovulation in cycling
    rats. Toxicol Appl Pharmacol, 128: 216-223.

    Davis BJ, Weaver R, Gaines LJ, & Heindel JJ (1994b) Mono-(2-
    ethylhexyl)phthalate suppresses estradiol production independent
    of FSH-cAMP stimulation in rat granulosa cells. Toxicol Appl
    Pharmacol, 128: 224-228.

    DeFoe DL, Holcombe GW, Hammermeister DE, & Biesinger KE (1990)
    Solubility and toxicity of eight phthalate esters to four aquatic
    organisms. Environ Toxicol Chem, 9: 623-636.

    DeVault DS (1985) Contaminants in fish from Great Lakes harbors
    and tributary mouths. Arch Environ Contam Toxicol, 14(5):
    587-594.

    Di Toro DM, Zarba CS, Hansen DJ, Berry WJ, Swartz RC, Cowan CE,
    Pavlou SP, Allen HE, Thomas NA, & Paquin PR (1991) Technical
    basis for establishing sediment quality criteria for nonionic
    organic chemicals using equilibrium partitioning. Environ Toxicol
    Chem, 10: 1541-1583.

    Eaton RW & Ribbons DW (1982) Metabolism of dibutyl phthalate and
    phthalate by  Micrococcus sp. strain 12B. J Bacteriol, 151:
    48-57.

    EEC (European Economic Community) (1987) Council directive of 25
    April 1983 on the approximation of the laws of the Member States
    relating to materials and articles made of regenerated cellulose
    film intended to come into contact with foodstuffs.

    EG&G Bionomics (1983a) Acute toxicity of fourteen phthalate
    esters to fathead minnows ( Pimephales promelas). Penascola,
    Florida, EG&G Bionomics (Toxicity test report submitted to
    Chemical Manufacturers Association, Washington, DC).

    EG&G Bionomics (1983b) Acute toxicity of thirteen phthalate
    esters to fathead minnows ( Pimephales promelas) under flow-
    through conditions. Penascola, Florida, EG&G Bionomics (Toxicity
    test report submitted to Chemical Manufacturers Association,
    Washington, DC).

    EG&G Bionomics (1983c) Acute toxicity of thirteen phthalate
    esters to bluegill ( Lepomis macrochirus). Penascola, Florida,
    EG&G Bionomics (Toxicity test report submitted to Chemical
    Manufacturers Association, Washington, DC).

    EG&G Bionomics (1983d) Acute toxicity of fourteen phthalate
    esters to rainbow trout ( Salmo gairdneri) under flow-through
    conditions. Penascola, Florida, EG&G Bionomics (Toxicity test
    report submitted to Chemical Manufacturers Association,
    Washington, DC).

    EG&G Bionomics (1984a) Acute toxicity of twelve phthalate esters
    to mysid shrimp ( Mysidopsis bahia). Penascola, Florida, EG&G
    Bionomics (Toxicity test report submitted to Chemical
    Manufacturers Association, Washington, DC).

    EG&G Bionomics (1984b) Acute toxicity of twelve phthalate esters
    to  Paratanytarsus parthenogenica. Penascola, Florida, EG&G
    Bionomics (Toxicity test report submitted to Chemical
    Manufacturers Association, Washington, DC).

    Eglinton G, Simoneit BRT, & Zoro JA (1975)  he recognition of
    organic pollutants in aquatic sediments. Proc R Soc (Lond),
    B189: 415-442.

    Ehrhardt M & Derenbach J (1980) Phthalate esters in the Kiel
    Bight. Mar Chem, 8: 339-346.

    Eisenreich SJ, Looney BB, & Thornton JD (1981) Airborne organic
    contaminants in the Great Lakes ecosystem. Environ Sci Technol,
    15: 30-38.

    Elsisi AE, Carter DE, & Spies IG (1989) Dermal absorption of
    phthalate diesters in rats. Fundam Appl Toxicol, 12(1): 70-77.

    Ema M, Amano H, Itami T, & Kawasaki H (1993) Teratogenic
    evaluation of di-n-butyl phthalate in rats.  Toxicol Lett,
    69(2): 197-203.

    Ema M, Amano H, & Ogawa Y (1994) Characterization of the
    developmental toxicity of di-n-butyl phthalate in rats.
    Toxicology, 86(3): 163-174.

    Fallon ME & Horvath FJ (1985) Preliminary assessment of
    contaminants in soft sediments of the Detroit River. J Great
    Lakes Res, 11: 373-378.

    Fatoki OS & Vernon F (1990) Phthalate esters in rivers of the
    Greater Manchester area, U.K. Sci Total Environ, 95: 227-232.

    Fishbein L & Albro PW (1972) Chromatographic and biological
    aspects of the phthalate esters. J Chromatogr, 70: 365-412.

    Fischer J, Ventura K, Prokes B, & Janderaa P (1993) Methods for
    determination of plasticizers in industrial emissions.
    Chromatographia, 37(1/2): 47-50.

    Florin I, Rutberg L, Curvall M, & Enzell CR (1980) Screening of
    tobacco smoke constituents for mutagenicity using the Ames' test.
    Toxicology, 15: 219-232.

    Foster PMD, Lake BG, Thomas LV, Cook MW, & Gangolli D (1981)
    Studies on the testicular effects and zinc excretion produced by
    various isomers of monobutyl-o-phthalate in the rat. Chem-Biol
    Interact, 34: 233-238.

    Foster PM, Cook MW, Thomas LV, Walters DG, & Gangolli SD (1982)
    Differences in urinary metabolic profile from di-n-butyl
    phthalate-treated rats and hamsters:  A possible explanation for
    species differences in susceptibility to testicular atrophy. Drug
    Metab Dispos, 11(1): 59-61.

    Fredricsson B, Mœller L, Pousette Å, & Westerholm R (1993) Human
    sperm motility is affected by plasticizers and diesel particle
    extracts. Pharmacol Toxicol, 72(2): 128-133.

    Fukuoka M, Tanimoto T, Zhou Y, Kawasaki N, Tanaka A, Ikemoto I, &
    Machida T (1989)  Mechanism of testicular atrophy induced by
    di-n-butyl phthalate in rats. Part 1. J Appl Toxicol, 9(4):
    277-283.

    Fukuoka M, Zhou Y, Tanaka A, Ikemoto I, & Machida T (1990)
    Mechanism of testicular atrophy induced by di-n-butyl phthalate
    in rats. Part 2. The effects on some testicular enzymes. J Appl
    Toxicol, 10(4):
    285-293.

    Fukuoka M, Kobayashi T, Zhou Y, & Hayakawa T (1993) Mechanisms of
    testicular atrophy induced by di- n-butyl phthalate in rats.
    Part 4. Changes in the activity of succinate dehydrogenase and
    the levels of transferrin and ferritin in the Sertoli and germ
    cells. J Appl Toxicol, 13(4): 241-246.

    Furtmann K (1994) Phthalate in surface water - a method for
    routine trace level analysis. Fresenius J Anal Chem, 348:
    291-296.

    Giam CS & Atlas E (1980) Accumulation of phthalate ester
    plasticizers in Lake Constance sediments.  Nat.wiss, 67:
    508-510.

    Giam CS, Chan HS, Neff GS, & Atlas EL (1978) Phthalate ester
    plasticizers: a new class of marine pollutant. Science, 199:
    419-421.

    Giam CS, Atlas E, & Chan HS (1980) Phthalate esters, PCB and DDT
    residues in the Gulf of Mexico atmosphere. Atmos Environ, 14:
    65-69.

    Gilioli R, Bulgherain C, Terrana T, Filippini G, Massette N, &
    Boeri R (1978) Horizontal and longitudinal study of a population
    employed in the production of phthalates. Med Lav, 69: 620-631.

    Glass GE, Strachan WMI, Willford WA, Armstrong FAI, Kaiser KLE, &
    Lutz A (1977) Organic contaminants. In: The waters of Lake Huron
    and Lake Superior. Volume III, Part B: Lake Superior.  Detroit,
    Michigan, The Upper Lakes Reference Group, pp 417-429, 499-502
    (Report to the International Joint Commission) (EPA 600/J-77-
    042).

    Golder Associates (1987) Testing of specific organic compounds in
    soils in background in urban areas, Port Credit and
    Oakville/Burlington, Ontario. Mississauga, Ontario, Canada,
    Golder Associates (Working paper to Shell Canada Limited and
    Texaco Canada Limited).

    Golder Associates, SENES Consultants Limited and CanTox (1987)
    Summary interpretation of the influence of soil conditions on the
    decommissioning of the Oakville refinery site. Mississauga,
    Ontario, Canada, Golder Associates (Working paper No. 4 to Shell
    Canada Limited).

    Government of Canada (1993) Canadian Environmental Protection Act
    - Priority substances list. Supporting document: Dibutyl
    phthalate. Ottawa, Ontario, Environment Canada.

    Government of Canada (1994) Canadian Environmental Protection Act
    - Priority substances list. Assessment report: Dibutyl phthalate.
    Ottawa, Ontario, Environment Canada and Health Canada.

    Graham PR (1973) Phthalate ester plasticizers - Why and how they
    are used. Environ Health Perspect, 3: 3-12.

    Gray TJB & Butterworth KR (1980) Testicular atrophy produced by
    phthalate esters. Arch Toxicol, 4(Suppl): 452-455.

    Gray TJB, Rowland IR, Foster PMD, & Gangolli SD (1982) Species
    differences in the testicular toxicity of phthalate esters.
    Toxicol Lett, 11: 141-147.

    Gray LE Jr, Laskey JW, Ostby J, & Ferrell J (1983a) The effects
    of dibutyl phthlate on the reproductive tract of the male and
    female rat and hamster. Toxicologist, 3: 22 (Abstract No. 87).

    Gray TJB, Lake B, Beamand A, Foster JR, & Gangolli SD (1983b)
    Peroxisomal effects of phthlate esters in primary cultures of rat
    hepatocytes. Toxicology, 28: 167-179.

    Guardiola J, Ventura F, & Rivera J (1989) Occurrence of
    industrial organic pollution in a groundwater supply:  screening,
    monitoring and evaluation of treatment processes. Water Supply,
    7: 11-16.

    Hamano Y, Kuwano A, Inoue K, Oda Y, Yamamoto H, Mitsuda B, &
    Kunita N (1977) Studies on toxicity of phthalic acid esters.
    First report: Teratogenic effects in mice administered orally.
    Osaka-furitsu Koshu Esei kenkyusho Kenkyu Hokoka Shokukhim Eisei
    Hen, 8: 29-33.

    Hardin BD, Schuler RL, Burg JR, Booth GM, Hazelden KP, MacKenzie
    KM, Piccirillo VJ, & Smith KN (1987) Evaluation of 60 chemicals
    in a preliminary developmental toxicity test. Teratogen
    Carcinogen Mutagen, 7: 29-48.

    Hardwick RC, Cole RA, & Fyfield TP (1984) Injury to and death of
    cabbage ( Brassica oleracea) seedlings caused by vapours of di
    butyl phthalate emitted from certain plastics. Ann Appl Biol,
    105: 97-105.

    Hazleton Biotechnologies Company (1986) Mutagenicity of 1C in a
    mouse lymphoma mutation assay.  Final Report (Genetics assay No.
    7156 - HB Project No. 20989). Kensington, Maryland, Hazleton
    Biotechnologies Company (Submitted by Chemical Manufacturers
    Association to US Environmental Protection Agency) (EPA Document
    No. 40-8626225; microfiche No. OTS0510527).

    Heindel JJ, Gulati DK, Mounce RC, Russell SR, & Lamb JC IV (1989)
    Reproductive toxicity of three phthalic acid esters in a
    continuous breeding protocol. Fundam Appl Toxicol, 12(3):
    508-518.

    Herring R & Bering CL (1988) Effects of phthalate esters on plant
    seedlings and reversal by a soil microorganism. Bull Environ
    Contam Toxicol, 40(4): 626-632.

    Hibino M, Matsuda H, Sato T, Ose Y, Nagase H, & Kito H (1992)
    Generation of mutagenicity by ozonation of humic substances'
    components. Sci Total Environ, 16: 1-13.

    Hill EF, Heath RG, Spann JW, & Williams JD (1975) Lethal dietary
    toxicities of environmental pollutants to birds. Washington, DC,
    US Department of the Interior, Fish and Wildlife Service (Special
    Scientific Report - Wildlife No. 191).

    Hites RA & Budde WL (1991) EPA's analytical methods for water:
    the next generation. Environ Sci Technol, 25(6): 998-1006.

    Ho CT, Lee KN, & Jin QZ (1983) Isolation and identification of
    volatile flavor compounds in fried bacon. J Agric Food Chem,
    31: 336-342.

    Hoff RM & Chan KW (1987) Measurement of polycyclic aromatic
    hydrocarbons in air along the Niagara river. Environ Sci Technol,
    21(6): 556-561.

    Howard PH ed. (1989) Handbook of environmental fate and exposure
    data for organic chemicals. Chelsea, Michigan, Lewis Publishers
    Inc., 574 pp.

    Howard PH, Boethling RS, Jarvis WF, Meylan WM, & Michalenko EM
    (1991) Handbook of environmental degradation rates. Chelsea,
    Michigan, Lewis Publishers Inc.

    Hrudey SE, Sergy GA, & Thackeray T (1976) Toxicity of oil sands
    plant wastewaters and associated organic contaminants. In: Water
    pollution research in Canada: Proceedings of the 11th Canadian
    Symposium. Ottawa, Ontario, Water Pollution Research Canada, pp
    34-45.

    HSE (Health and Safety Executive) (1986) Review of the toxicity
    of the esters of o-phthalic acid (phthalate esters). Sudbury,
    United Kingdom, HSE Books, 183 pp (Toxicity Review 14).

    Hudson RA & Bagshaw JC (1978) Toxicity of di- n-butyl phthalate
    for developing larvae of the brine shrimp,  Artemia salina. Fed
    Proc, 37: 1702 (Abstract No. 2393).

    Hudson RA, Austerberry CF, & Bagshaw JC (1981) Phthalate ester
    hydrolases and phthalate ester toxicity in synchronously
    developing larvae of the brine shrimp ( Artemia). Life Sci,
    29: 1865-1872.

    Husain SL (1975) Dibutyl phthalate sensitivity. Contact
    Dermatitis, 1: 395.

    Ikemoto I (1985) Experimental studies on testicular damage
    induced by phthalic acid esters. Tokyo Jikeikai Med J, 100:
    1115-1127.

    Ikemoto I, Kotera S, Katsurai K, Inaba Y, Machida T, & Tanaka A
    (1983) Experimental testicular damage induced by dibutyl
    phthalate and monobutyl phthalate. Jpn J Fertil Steril, 28(7):
    159-165.

    Inman JC, Strachan SD, Sommers LE, & Nelson DW (1984) The
    decomposition of phthalate esters in soil. J Environ Sci Health,
    B19: 245-257.

    IPCS (1992) Environmental health criteria 131: Diethylhexyl
    phthalate. Geneva, World Health Organization, International
    Programme on Chemical Safety, 141 pp.

    IPCS (1993) Environmental health criteria 170: Assessing human
    health risks of chemicals: Derivation of guidance values for
    health-based exposure limits. Geneva, World Health Organization,
    International Programme on Chemical Safety, 73 pp.

    Ishida M, Suyama K, & Adachi S (1980) Background contamination by
    phthalates commonly encountered in the chromatographic analysis
    of lipid samples. J Chromatogr, 189: 421-424.

    Ishida M, Suyama K, & Adachi S (1981) Occurrence of dibutyl and
    di(2-ethylhexyl) phthalate in chicken eggs. J Agric Food Chem,
    29: 72-74.

    Ishidate M & Odashima S (1977) Chromosome tests with 134
    compounds on Chinese hamster cells  in vitro - a screening for
    chemical carcinogens. Mutat Res, 48(3/4): 337-354.

    Ito S, Takeda H, Kobayashi A, Sakurai H, Tada Y, Aoki G, Hosogai
    T, Yamanaka T, & Ishiwata H (1993) A simple and rapid method for
    determination of n-dibutylphthalate in imported vodka by FID-GC
    and GC/MS. J Food Hyg Soc Jpn, 34(3): 254-256.

    Iturbe R, Elefsiniotis P, & Moreno G (1991) Efficiency of a
    phthalate ester in an activated sludge system. Environ Technol,
    19(9): 783-796.

    Johnson BT & Lulves W (1975) Biodegradation in di- n-butyl
    phthalate and di-(2-ethylhexyl) phthalate in freshwater
    hydrosoil. J Fish Res Board Can, 32(3): 333.

    Johnson BT, Stalling DL, Hogan JW, & Schoettger RA (1977)
    Dynamics of phthalate acid esters in aquatic organisms. In:
    Suffet IH ed. Fate of pollutants in the air and water
    environment. Part 2. New York, Wiley Interscience, pp 283-300
    (Advances in Environmental Science and Technology, Volume 8).

    Johnson BT, Heitkamp MA, & Jones JR (1984) Environmental and
    chemical factors influencing the biodegradation of phthalic acid
    esters in freshwater sediments. Environ Pollut, B8: 101-118.

    JPIF (1995) Production of plastics. Plastics, 46: 104.

    Kaneshima H, Yamaguchi T, Okui T, & Naitoh M (1978) Studies on
    the effects of phthalate esters on the biological system. Part 2:
     In vitro metabolism and biliary excretion of phthalate esters
    in rats. Bull Environ Contam Toxicol, 19: 502-509.

    Kawano M (1980a) Toxicological studies on phthlate esters. 2.
    Metabolism, accumulation and excretion of phthlate esters in
    rats. Jpn J Hyg, 35: 693-701.

    Kawano M (1980b) Toxicological studies on phthlate esters. 1.
    Inhalation effects of dibutyl phthalate (DBP) on rats. Jpn J Hyg,
    35: 684-692.

    Kawashima Y, Hanioka N, Matsumura M, & Kozuka H (1983) Induction
    of microsomal stearoyl-CoA desaturation by the administration of
    various peroxisome proliferators. Biochim Biophys Acta, 752:
    259-264.

    Keith LH, Garrison AW, Allen FR, Carter MH, Floyd TL, Pope JD, &
    Thruston AD (1976) Identification of organic compounds in
    drinking water from thirteen U.S. cities.  In: Keith LH ed. 
    Identification and analysis of organic pollutants in water. Ann
    Arbor, Michigan, Ann Arbor Science Publishers Inc., pp 329-373.

    Khaliq MA, Alam MS, & Srivastava SP (1992) Implications of
    physico-chemical factors on the migration of phthalate esters
    from tubing commonly used for oral/nasal feeding. Bull Environ
    Contam Toxicol, 48: 572-578.

    Klöpfer W, Kaufmann G, Rippen G, & Poremski HJ (1982) A
    laboratory method for testing the volatility from aqueous
    solution: first results and comparison with theory. Ecotoxicol
    Environ Saf, 6: 545-559.

    Kohli J, Ryan JF, & Afghan BK (1989) Phthalate esters in the
    aquatic environment. In: Chau BK & Chau ASY ed. Analysis of trace
    organics in the aquatic environment. Boca Raton, Florida, CRC
    Press Inc., pp 243-281.

    Kördel W & Müller J (1992) Occurrence of phthalic acid esters in
    soil and plants nearby phthalate-emitting sources and by sewage
    sludge treatment. Schmallenberg-Grafshaft, Germany, Fraunhofer
    Institute for Environmental Chemistry and Ecotoxicology.

    Korhonen A, Hemminki K, & Vanio H (1983) Embryotoxic effects of
    phthalic-acid derivatives phosphates and aromatic oils used in
    the manufacturing of rubber on 3 day chicken embryos. Drug Chem
    Toxicol, 6(2): 191-208.

    Kozumbo WJ, Kroll R, & Rubin RJ (1982) Assessment of the
    mutagenicity of phthalate esters. Environ Health Perspect, 45:
    103-109.

    Krauskopf LG (1973) Studies of the toxicity of phthalates via
    ingestion. Environ Health Perspect, 3: 61-72.

    Kühn R & Pattard M (1990) Results of the harmful effects of water
    pollutants to green algae ( Scenedesmus subspicatus) in the
    cell multiplication inhibition test. Water Res, 24(1): 31-38.

    Kühn R, Pattard M, Pernak, KD, & Winter A (1989) Results of the
    harmful effects of water pollutants to  Daphnia magna in the 21
    day reproduction test. Water Res, 23(4): 501-510.

    Kurane R, Suzuki T, & Takahara Y (1979a) Microbial population and
    identification of phthalate ester-utilizing microorganisms. Agric
    Biol Chem, 43: 907-917.

    Kurane R, Suzuki T, & Takara Y (1979b) Microbial degradation of
    phthalic esters. Agric Biol Chem, 43(3): 421-427.

    Kveseth NJ (1980) Chlorinated hydrocarbons in sewage sludge from
    a plant in Oslo.  Nord Vet Med, 32: 341-347.

    Lake BG, Phillips JC, Linnell JC, & Gangolli SD (1977) The
     in vitro hydrolysis of some phthalate diesters by hepatic and
    intestinal preparations from various species. Toxicol Appl
    Pharmacol, 39: 239-248.

    Lake BG, Cook WM, Worrell NR, Cunninghame ME, Evans JG, Price RJ,
    Young PJ, & Carpanini FMB (1991) Dose-response relationships for
    induction of hepatic peroxisome proliferation and testicular
    atrophy by phthalate esters in the rat. Hum Exp Toxicol, 10:
    67-68 (Abstract).

    Lamb JC, Chapin RE, Teague J, Lawton AD, & Reel JR (1987)
    Reproductive effects of four phthalic acid esters in the mouse.
    Toxicol Appl Pharmacol, 88(2): 255-269.

    Laughlin RB Jr, Neff JM, Hrung YC, Goodwin TC, & Giam CS (1978)
    The effects of three phthalate esters on the larval development
    of the grass shrimp  Palaemonetes pugio (Holthuis). Water Air
    Soil Pollut, 9: 323-336.

    Law RJ, Fileman TW, & Matthiessen P (1991) Phthalate esters and
    other industrial organic chemicals in the North and Irish Seas.
    Water Sci Technol, 24: 127-134.

    Lawrence WH, Malik M, Turner JE, Singh AR, & Autian J (1975) A
    toxicological investigation of some acute, short-term and chronic
    effects of administering di(2-ethylhexyl) phthalate (DEHP) and
    other phthalate esters. Environ Res, 9: 1-11.

    Lehman AJ (1955) Insect repellents. Assoc Food Drug Off, 19:
    87-99.

    Leibowitz JN, Sarmiento R, Dugar SM, & Ethridge MW (1995)
    Determination of six common phthalate plasticisers in grain
    neutral spirits and vodka. J Assoc Off Anal Chem Int 78(3):
    730-735.

    Lesage S (1991) Characterization of groundwater contaminants
    using dynamic thermal stripping and adsorption/thermal
    desorption-GC-MS. Fresenius J Anal Chem, 339: 516-527.

    Lewis RJ (1991) Reproductively active chemicals: A reference
    guide. New York, Van Nostrand Reinhold Co.

    Leyder F & Boulanger P (1983) Ultraviolet absorption, aqueous
    solubility, and octanol-water partition for several phthalates.
    Bull Environ Contam Toxicol, 30(2): 152-157.

    Ligocki MP & Pankow JF (1985) Assessment of absorption/solvent
    extraction with polyurethane foam and adsorption/thermal
    desorption with Tenax-GC for the collection and analysis of
    ambient organic vapours. Anal Chem, 57: 1138-1144.

    Ligocki MP, Leuenberger C, & Pankow JF (1985) Trace organic
    compounds in rain-II. Gas scavenging of neutral organic
    compounds. Atmos Environ, 19: 1609-1617.

    Lindén E, Bengtsson BE, Svanberg O, & Sundström G (1979) The
    acute toxicity of 78 chemicals and pesticide formulations against
    two brackish water organisms, the bleak ( Alburnus alburnus)
    and the harpacticoid ( Nitocra spinipes). Chemosphere,
    8(11/12): 843-851.

    Litton Bionetics Inc. (1985) Evaluation of 1C in the  in vitro
    transformation of Balb/3T3 cellsaAssay.  Final report (Genetics
    assay No. 7156 - LBI Project No. 20992). Kensington, Maryland,
    Litton Bionetics Inc. (Submitted by Chemical Manufacturers
    Association to US Environmental Protection Agency (EPA document
    No. 40+8526206; microfiche No. OTS0509537).

    Lokke H & Bro-Rasmussen F (1981) Studies of mobility of di-iso-
    butyl phthalate (DiBP), di-n-butyl phthalate (DBP), and di-(2-
    ethyl hexyl) phthalate (DEHP) by plant foliage treatment in a
    closed terrestrial simulation chamber. Chemosphere, 10(11/12):
    1223-1235.

    Lunde G, Gether J, GjØs N, & StØbet Lande MB (1977) Organic
    micropollutants in precipitation in Norway. Atmos Environ, 11:
    1007-1014.

    LWA (1993) Report on the quality of the Rhine.  Düsseldorf,
    Northrhine-Westfalia Office for Water and Waste.

    McCarthy JF & Whitmore DK (1985) Chronic toxicity of di-n-butyl
    and di-n-octyl phthalate to  Daphnia magna and the fathead
    minnow. Environ Toxicol Chem, 4: 167-179.

    McKone TE & Layton DW (1986) Exposure and risk assessment of
    toxic waste in a multimedia context. 79th Annual Meeting of the
    Air Pollution Control Association, Minneapolis, Minnesota, 22-27
    June 1986. Pittsburgh, Pennsylvania, Air Pollution Control
    Association, vol 1, 16 pp.

    MAFF (Ministry of Agriculture, Fisheries and Food, United
    Kingdom) (1987)  Survey of plasticiser levels in food contact
    materials and in foods. The 21st report of the Steering Group on
    Food Surveillance/The Working Party on Chemical Contaminants from
    Food Contact Materials / Sub Group on Plasticisers. London, Her
    Majesty's Stationery Office, 105 pp (Food  Surveillance Paper
    No. 21).

    MAFF (Ministry of Agriculture, Fisheries and Food, United
    Kingdom) (1990) Plasticizers:  Continuing surveillance. The 31st
    report of the Steering Group on Food Surveillance/The Working
    Party on Chemical Contaminants from Food Contact Materials / Sub
    Group on Plasticisers. London, Her Majesty's Stationery Office,
    53 pp (Food Surveillance Paper No. 30).

    Malisch R, Schulte E, & Acker L (1981) [Organochlorine
    pesticides, polychlorinated biphenyls and phthalate in Rhine and
    Neckar sediments.] Chem Ztg, 105: 187-194 (in German).

    Mathur SP (1974) Phthalate esters in the environment: pollutants
    or natural products? J Environ Qual, 3(3): 189-197.

    Matsuda K & Schnitzer M (1971) Reactions between fulvic acid, a
    soil humic material, and dialkyl phthalates. Bull Environ Contam
    Toxicol, 6(3): 200-204.

    Mayer FL Jr & Sanders HO (1973) Toxicology of phthalic acid
    esters in aquatic organisms. Environ Health Perspect, 3:
    153-157.

    Mayer FL & and Ellersieck MR (1986) Manual of acute toxicity:
    interpretation and data base for 410 chemicals and 66 species of
    freshwater animals. Washington, DC, US Department of the
    Interior, Fish and Wildlife Service (Resource Publication 160;
    NTIS Publication No. PB86-239878).

    Melin C & Egnéus H (1983) Effects of di- n-octyl phthalate on
    growth and photosynthesis in algae and on isolated organelles
    from higher plants. Physiol Plant, 59: 461-466.

    Men'shikova TA (1971) Hygienic evaluation of dibutyl phthlate in
    relation to the use of polymer finishes in shipboard living
    quarters. Hyg Sanit, 36: 349-353.

    Mes J, Coffin DE, & Campbell DS (1974) Di-n-butyl- and di-2-
    ethylhexyl phthalate in human adipose tissue. Bull Environ Contam
    Toxicol, 12(6): 721-725.

    Milkov LE, Aldyrova MV, Popova TB, Lopukhova KA, Makarenko YL,
    Malyar LM, & Shakhova TK  (1973) Health status of workers exposed
    to phthalate plasticizers in the manufacture of artificial
    leather and films based on PVC resins. Environ Health Perspect,
    3: 175-178.

    Millar DJ & Hannay JW (1986) Phytotoxicity of phthalate
    plasticisers. 2. Site and mode of action. J Exp Bot, 37(179):
    898-908.

    MOE (Ontario Ministry of the Environment) (1984) Drinking water
    survey of selected municipalities in the Niagara area and Lake
    Ontario. Ottawa, Ontario, Ontario Ministry of the Environment.

    Montgomery JH & Welkom LM (1990) Groundwater chemicals desk
    reference. Chelsea, Michigan, Lewis Publishers Inc.

    Murakami K, Nishiyama K, & Higuti T (1986a) Toxicity of dibutyl
    phthalate and its metabolites in rats.  Jpn J Hyg, 41(4):
    775-780.

    Murakami K, Nishiyama K, & Higuti T (1986b) Mitochrondrial effect
    of orally administered dibutyl phthalate in rats. Jpn J Hyg,
    41(4): 769-774.

    Nagy KA (1987) Field metabolic rate and food requirement scaling
    in mammals and birds. Ecol Monogr, 57(2): 111-128.

    NAQUADAT (1993) National water quality data bank. Ottawa,
    Ontario, Environment Canada, Inland Waters Directorate, Water
    Quality Branch.

    Nerin C, Cacho J, & Gancedo P (1993) Plasticizers from printing
    inks in a selection of food packagings and their migration to
    food. Food Addit Contam, 10(4): 453-460.

    Neuhauser EF, Loehr RC, Malecki MR, Milligan DL, & Durkin PR
    (1985) The toxicity of selected organic chemicals to the
    earthworm  Eisenia  fetida. J Environ Qual, 14(3): 383-388.

    Neuhauser EF, Loehr RC, & Malecki MR (1986)  Contact and
    artificial soil tests using earthworms to evaluate the impact of
    wastes in soil. In: Petros JK, Lacy WJ, & Conway RA ed. Fourth
    Symposium on Hazardous and Industrial Solid Waste Testing.
    Philadelphia, Pennsylvania, American Society for Testing and
    Materials, pp 192-203 (ASTM STP 886).

    Niagara River Data Interpretation Group (1990) Joint evaluation
    of upstream/downstream Niagara River monitoring data, 1988-1989.
    Prepared by Data Interpretation Group, River Monitoring
    Committee. Joint publication of Environment Canada, US
    Environmental Protection Agency, Ontario Ministry of the
    Environment, and New York State Department of Environmental
    Conservation.

    Nikonorow M, Mazur H, & Piekacz H (1973) Effect of orally
    administered plasticizers and polyvinyl chloride stabilizers in
    the rat. Toxicol Appl Pharmacol, 26: 253-259.

    NIOSH (1976) Health hazard evaluation determination report
    No. 74-120-260: Goodyear Tire and Rubber Company, Gadsden,
    Alabama. Cincinnati, Ohio, National Institute for Occupational
    Safety and Health (NTIS Publication No. PB89-166078).

    NIOSH (1982)  Health hazard evaluation report No. HETA-81-277-
    1089: Indiana Army Ammunition Plant, Charlestown, Indiana.
    Cincinnati, Ohio, National Institute for Occupational Safety and
    Health (NTIS Publication No. PB83-198697).

    NIOSH (1987) Health hazard evaluation report No. MHETA 86-191-
    1836: West Virginia Department of Highways, Charleston, West
    Virginia. Cincinnati, Ohio, National Institute for Occupational
    Safety and Health (NTIS Publication No. PB88-162698).

    NIOSH (1989) Health hazard evaluation report No. MHETA 88-214-
    1952: Flying W Plastics Company, Glenville, West Virginia.
    Cincinnati, Ohio, National Institute for Occupational Safety and
    Health (NTIS Publication No. PB89-230544).

    NTP (1984) Di(n-butyl) phthalate: Reproduction and fertility
    assessment in CD-1 mice when administered in the feed. Research
    Triangle Park, North Carolina, US Department of Health and Human
    Services, National Toxicology Program, 100 pp (NTIS Publication
    No. PB85-144798).

    NTP(1991) Final report on the reproductive toxicity of di-n-butyl
    phthalate (CAS No. 84-74-2) in Sprague-Dawley rats. Research
    Triangle Park, North Carolina, US Department of Health and Human
    Services, National Toxicology Program (Report No. T-0035C; NTIS
    Publication No. PB92-111996).

    NTP (1994a) Chairperson's report. Pathology Working Group (PWG)
    review - Dibutylphthalate (C62022B) and (C62022): A 13-week
    subchronic toxicity study (Phases II and III) administered by
    dose feed in F344 rats and B6C3F1 mice conducted at Battelle
    Columbus. Research Triangle Park, North Carolina, US Department
    of Health and Human Services, National Toxicology
    Program/Environmental Toxicology Program.

    NTP (1994b) Prechronic study, dibutyl phthalate diet study in
    mice. Research Triangle Park, North Carolina, US Department of
    Health and Human Services, National Toxicology
    Program/Environmental Toxicology Program.

    NTP (1995) NTP technical report on toxicity studies of dibutyl
    phthalate (CAS No. 84-74-2) aministered in feed to F344/N rats
    and B6C3F1 mice. Research Triangle Park, North Carolina, US
    Department of Health and Human Services, National Toxicology
    Program (Toxicity Series No. 30).

    O'Connor OA, Rivera MD, & Young LY (1989) Toxicity and
    biodegradation of phthalic acid esters under methanogenic
    conditions. Environ Toxicol Chem, 8: 569-574.

    Ogura N, Ambe Y, Ogura K, Ishiwatari R, Mizutani T, Satoh Y,
    Matsushima H, Katase T, Ochiai M, Tadokoro K, Takada T, Sugihara
    K,  Matsumoto G, Nakamoto N, Funakoshi M, & Hanya T (1975)
    Chemical composition of organic compounds present in water of the
    Tamagawa River.  Jpn J Limnol, 36: 23-30.

    Oishi S & Hiraga K (1980a) Testicular atrophy induced by phthalic
    acid esters. Effect on testosterone and zinc concentrations.
    Toxicol Appl Pharmacol, 53: 35-41.

    Oishi S & Hiraga K (1980b) Effect of phthalic acid esters on
    mouse testes. Toxicol Lett, 5: 413-416.

    Oishi S & Hiraga K (1980c) Testicular atrophy induced by phthalic
    acid monoesters: Effects of zinc and testosterone concentrations.
    Toxicology, 15: 197-202.

    Oishi S & Hiraga K (1980d) Effects of phthalic acid monoesters on
    mouse testes. Toxicol Lett, 6: 239-242.

    Oishi S & Hiraga K (1982) Effects of monoesters of o-phthalic
    acid on serum lipid composition of rats. Toxicol Lett, 14:
    79-84.

    Ota H, Takashima K, Takashima Y, Onda H, Kodama H, & Yamada N
    (1973)  Biological effects of phthalate esters. (I)
    Histopathological findings from experiments in mice. Nippon
    Byorigakkai Kaishi, 62: 119-120.

    Ota H, Onda H, Kodama H, & Yamada N (1974) Histopathological
    studies on the effect of phthalic acid esters on the biological
    system of mice. Nippon Eiseigaku Kaishi, 29: 519-524.

    Otson R & Benoit FM (1985) Surveys of selected organics in
    residential air. In: Walkinshaw DS ed.  Indoor air quality in
    cold climates:  An Air Pollution Control Association speciality
    conference. Pittsburgh, Pennsylvania, Air Pollution Control
    Association, pp 224-236.

    Overcash MR, Weber JB, & Miles ML (1982) Behavior of organic
    priority pollutants in the terrestrial system: di-n-butyl
    phthalate ester, toluene, and 2,4-dinitrophenol. Raleign, North
    Carolina, University of North Carolina, Water Resources Research
    Institute (Report No. 171).

    PACE (Petroleum Association for the Conservation of the Canadian
    Environment) (1985) Evaluation of the variability of trace
    organic substances in petroleum refinery effluents. Ottawa,
    Ontario, Canadian Petroleum Products Institute (PACE Report
    No. 85-6).

    Page BD & Lacroix GM (1992) Studies into the transfer and
    migration of phthalate esters from aluminum foil-paper laminates
    to butter and margarine. Food Addit Contam, 9(3): 197-212.

    Page BD & Lacroix GM (1995) The occurrence of phthalate ester and
    di-2-ethylhexyl adipate plasticizers in Canadian packaging and
    food sampled in 1985-1989: a survey. Food Addit Contam, 12(1): 
    129-151.

    Pankow JF, Ligocki MP, & Rosen ME (1988) Adsorption/thermal
    desorption with small cartridges for the determination of trace
    aqueous semivolatile organic compounds. Anal Chem, 60: 40-47.

    Parkman H & Remberger M (1995) Phthalates in Swedish sediments.
    Stockholm, Swedish Environmental Research Institute (IVL-Report).

    Peakall DB (1974) Effects of di-n-butyl and di-2-ethylhexyl
    phthalate on the eggs of ring doves. Bull Environ Contam Toxicol,
    12: 698-702.

    Peakall DB (1975) Phthalate esters: occurrence and biological
    effects.  Residue Rev, 54: 1-41.

    Perwak J, Goyer M, & Schimke G (1981) An exposure and risk
    assessment for phthalate esters: Di(2-ethylhexyl) phthalate, di-
    n-butyl phthalate, dimethyl phthalate, diethyl phthalate, di-n-
    octyl phthalate, butyl benzyl phthalate. Washington, DC, US
    Environmental Protection Agency, Office of Water Regulations and
    Standards (EPA-440/4-81-020;NTIS Publication No. PB85-211936).

    Peters JW & Cook RM (1973) Effect of phthalate esters on
    reproduction in rats. Environ Health Perspect, 3: 91-94.

    Peterson JC & Freeman DH (1984) Variations of phthalate ester
    concentrations in sediments from the Chester River, Maryland. Int
    J Environ Anal Chem, 18(4): 237-252.

    Pierce RC, Mathur SP, Williams DT, & Boddington MJ (1980)
    Phthalate esters in the aquatic environment. Ottawa, Ontario,
    National Research Council of Canada, Associate Committee on
    Scientific Criteria for Environmental Quality (Report No. NRCC
    17583).

    Poole CF & Vibberley DG (1977) Determination of DEHP ih human
    placenta. J Chromatogr, 132: 511.

    Radeva M & Dinoyeva S (1966) [The toxicity of the plasticiser
    dibutyl phthlate with oral administration to white rats.] Khig i
    Zdrav, 9: 510-515 (in Russian).

    Ray LE, Murray HE, & Giam CS (1983a) Analysis of water and
    sediment from the Nueces Estuary/Corpus Christi Bay (Texas) for
    selected organic pollutants. Chemosphere, 12(7/8): 1039-1045.

    Ray LE, Murray HE, & Giam CS  1983b) Organic pollutants in marine
    samples from Portland, Maine.  Chemosphere, 12(7/8): 1031-1038.

    RIWA (1991) Annual Report of the Working Party on the Rhine and
    the Maas.  Part A:  The Rhine.  Amsterdam, National Institute for
    Water and Waste (RIWA).

    RIWA (1992) Annual Report of the Working Party on the Rhine and
    the Maas.  Part A:  The Rhine.  Amsterdam, National Institute for
    Water and Waste (RIWA).

    Rogers IH & Hall KJ (1987) Chlorophenols and chlorinated
    hydrocarbons in starry flounder ( Platichthys stellatus) and
    contaminants in estuarine sediments near a large municipal
    outfall. Water Pollut Res J Can,  22(2): 197-210.

    Rogers IH, Birtell IK, & Kruzynski GM (1986) Organic extractables
    in municipal wastewater, Vancouver, British Columbia. Water
    Pollut Res J Can, 21(2): 187-204.

    Rowland IR, Cottrell RC, & Phillips JC (1977) Hydrolysis of
    phthalate esters by the gastro-intestinal contents of the rat.
    Food Cosmet Toxicol, 15: 17-21.

    Saito K, Nakazato M, Ishikawa F, Fujinuma K, Moriyasu T, Nagayama
    T, Kobayashi M, Shioda H, & Kamimura K (1993) [Determination of
    methyl isocyanate in wine and dibutyl phthalate in vodka.] Kenkyu
    Nempo Tokyo-Toritsu Eisei Kenkyusho, 44: 119-127 (in Japanese).

    Sajiki J (1975a) [Pathological observations of phthalate ester
    (di- n-butyl phthalate) for mice with orally administrations.]
    Chibaken Eiseikenkyusho Nenpo, 24: 57-63 (in Japanese).

    Sajiki J (1975b) Biological effects of phthalate ester (di- n-
    butyl phthalate) orally administered in mice.] Chibaken
    Eiseikenkyusho Nenpo, 24: 64-72 (in Japanese).

    Sandmeyer EE & Kirwin CJ (1981) Esters. In: Clayton GD & Clayton
    FE ed. Patty's industrial hygiene and toxicology. Volume 2A:
    Toxicology, 3rd rev ed. New York, John Wiley and Sons Inc., 
    pp 2345-2346.

    Sato H, Sato N, & Ichihara N (1975) Rec-assay application for
    mutagenicity testing of phthalic acid esters. Hokkaido-ritsu
    Eisei Kenkyushcho, 1975: 146-147.

    Schacht RA (1974) Pesticides in the Illinois waters of Lake
    Michigan. Washington, DC, US Environmental Protection Agency,
    Office of Research and Development (EPA 660/3-74-002).

    Schleyer R, Renner I, & Mühlhausen D (1991) [Immission load -
    Consequences for grounwater quality.] Bonn, Germany, Federal
    Ministry for Environment, Nature Conservation and Nuclear Safety
    (in German).

    Schouten MJ, Copius Peereboom JW, & Brinkman UA Th (1979) Liquid
    chromatographic analysis of phthalate esters in Dutch river
    water. Int J Environ Anal Chem, 7: 13-23.

    Schwartz HE, Anzion CJM, Van Vliets HPM, Copius Peerbooms JW, &
    UATh Brinkman (1979) Analysis of phthalate esters in sediments
    from Dutch rivers by means of high performance liquid
    chromatography. Int J Environ Anal Chem, 6: 133-144.

    Scott RC, Dugard PH, Ramsey JD, & Rhodes C (1987)  In vitro
    absorption of some o-phthalate diesters through human and rat
    skin. Environ Health Perspect, 74: 223-227.

    Seed JL (1982) Mutagenic activity of phthalate esters in
    bacterial liquid suspension assays. Environ Health Perspect,
    45: 111-114.

    Seth PK, Agarwal DK, & Agarwal S (1981) Effect of phthalic acid
    esters on drug metabolizing enzymes. Bull Environ Contam Toxicol,
    26: 764-768.

    Shahin MM & von Borstel RC (1977) Mutagenic and lethal effects of
    alpha-dibutyl phthalate and trichloroethylene in  Saccharomyces
     cerevisiae. Mutat Res, 48: 173-180.

    Shanker R, Ramakrishna C, & Seth PK (1985) Degradation of some
    phthalic acid esters in soil. Environ Pollut, A39: 1-7.

    Shea KP (1971) The new-car smell. Environment, 13(8): 2-9.

    Shea PJ, Weber JB, & Overcash MR (1982) Uptake and phytotoxicity
    of di- n-butyl phthalate in corn ( Zea mays). Bull Environ
    Contam Toxicol, 29: 153-158.

    Sheldon LS & Hites RA (1978) Organic compounds in the Delaware
    River. Environ Sci Technol, 12(10): 1188-1194.

    Sheldon LS & Hites RA (1979) Sources and movement of organic
    chemicals in the Delaware River. Environ Sci Technol, 13(5):
    574-579.

    Shelton DR, Boyd SA, & Tiedje JM (1984) Anaerobic biodegradation
    of phthalic acid esters in sludge. Environ Sci Technol, 18(2):
    93-97.

    Shibuya S (1979) Phthalic acid esters as one of the marker
    environmental pollutants - occurrence in the water and aquatic
    environment in Shizuoka Prefecture. Numazu Kogyo Kota Semmon
    Gakko Kenkyu Hokoku, 14: 63-72.

    Shiota K & Nishimura H (1982) Teratogenicity of di(2-ethylhexyl)
    phthalate (DEHP) and di-n-butyl phthalate (DBP) in mice. Environ
    Health Perspect, 45: 65-70.

    Shiota K, Chou MJ, & Nishimura H (1980) Embryotoxic effects of
    di(2-ethylhexyl)phthalate (DEHP) and di-n-butyl phthalate (DBP)
    in mice. Environ Res, 22: 245-253.

    Singh AR, Lawrence WH, & Autian J (1972) Teratogenicity of
    phthalate esters in rats. J Pharm Sci, 61(1): 51-55.

    Smith CC (1953) Toxicity of butyl stearate, dibutyl sebacate,
    dibutyl phthalate and methoxyethyl oleate.  Ind Hyg Occup Med,
    7: 310-318.

    Sneddon IB (1972) Dermatitis from dibutyl phthalate in an aerosol
    antiperspirant and deodorant.  Contact Dermatitis, 1972: 308.

    Spasovski M (1964) Experimental data to establish a maximum
    allowable concentration for dibutyl phthalate. Higiena, 7: 
    38-44.

    Springborn Bionomics (1984a) Toxicity of fourteen phthalate
    esters to the freshwater green algae,  Selenastrum
    capricornutum. Wareham, Massachusetts, Springborn Bionomics, 52
    pp (Toxicity test report submitted to the Chemical Manufacturers
    Association, Washington, DC).

    Springborn Bionomics (1984b) Chronic toxicity of fourteen
    phthalate esters to  Daphnia magna. Wareham, Massachusetts,
    Springborn Bionomics (Toxicity test report No. BW-84-5-1567
    submitted to the Chemical Manufacturers Association, Washington,
    DC).

    Srivastava SP, Srivastava S, Saxena DK, Chandra SV, & Seth PK
    (1990a) Testicular effects of di-n-butyl phthalate (DBP):
    Biochemical and histopathological alterations. Arch Toxicol,
    64(2): 148-152.

    Srivastava S, Singh GB, Srivastava SP, & Seth PK (1990b)
    Testicular toxicity of di-n-butyl phthalate in adult rats: Effect
    on marker enzymes of spermatogenesis. Ind J Exp Biol, 28(1): 
    67-70.

    Stalling DL, Hogan JW, & Johnson JL (1973) Phthalate ester
    residues - their metabolism and analysis in fish. Environ Health
    Perspect, 3: 159-173.

    Stanley JS (1986) Broad scan analysis of the FY82 national human
    adipose tissue survey specimens -  Volume I: Executive summary.
    Washington, DC, US Environmental Protection Agency,  Office of
    Toxic Substances (EPA 560/5-86-035).

    Streufert JM & Sanders HO (1977) Chronic effects of two phthalic
    acid esters on midge  Chironomus plumosus. TransMo Acad Sci,
    10/11: 297.

    Streufert JM, Jones JR, & Sanders HO (1980) Toxicity and
    biological effects of phthalate esters on midges ( Chironomus
     plumosus). TransMo Acad Sci, 14: 33-40.

    Sugatt RH, O'Grady DP, Banerjee S, Howard PH, & Gledhill WE
    (1984) Shake flask biodegradation of 14 commercial phthalate
    esters. Appl Environ Microbiol, 47: 601-606.

    Sugawara N (1974a) Toxic effect of a normal series of phthalate
    esters on the hatching of shrimp eggs. Toxicol Appl Pharmacol,
    30(1): 87-89.

    Sugawara N (1974b) Effect of phthalate esters on shrimp. Bull
    Environ Contam Toxicol, 12(4): 421-424.

    Sullivan KF, Atlas EL, & Giam CS (1982) Adsorption of phthalic
    acid esters from seawater. Environ Sci Technol, 16: 428-432.

    Swain WR (1978) Chlorinated organic residues in fish, water, and
    precipitation from the vicinity of Isle Royale, Lake Superior. J
    Great Lakes Res, 4: 398-407.

    Swain LG & Walton DG (1989) Report on the 1988 Fish Monitoring
    Program. British Columbia, Ministry of the Environment.

    Swartz RC, Schults DW, Ditsworth GR, DeBen WA, & Cole FA  (1985)
    Sediment toxicity, contamination, and macrobenthic communities
    near a large sewage outfall. In: Boyle TD ed. Validation and
    predictabolity of laboratory methods for assessing the fate and
    effects of contaminants in aquatic ecosystems. Philadelphia,
    Pennsylvania, American Society for Testing and Materials,
    pp 152-175 (ASTM STP 865).

    Tagatz ME, Deans CH, Moore JC, & Plaia GR (1983) Alterations in
    composition of field- and laboratory-developed estuarine benthic
    communities exposed to di-n-butyl phthalate. Aquatic Toxicol,
    3: 239-248.

    Tagatz ME, Plaia GR, & Deans CH  (1986) Toxicity of dibutyl
    phthalate-contaminated sediment to laboratory-and field-colonized
    estuarine benthic communities. Bull Environ Contam Toxicol,
    37(1): 141-149.

    Takahashi T & Tanaka A (1989) Biochemical studies on phthalic
    esters.V. Comparative studies on  in vitro hydrolysis of di-n-
    butyl phthalate isomers in rats. Arch Toxicol, 63: 72-74.

    Tanaka A, Matsumoto A, & Yamaha T (1978) Biochemical studies on
    phthalic esters. III. Metabolism of dibutyl phthalate (DBP) in
    animals. Toxicology, 9: 109-123.

    Tanino M, Ikemoto I, & Tanaka A (1987) Enzyme levels in rat
    testis damaged experimentally with dibutyl phthalate. Jikeikai
    Med J, 34: 245-252.

    Tarkpea M, Hansson M, & Samuelsson B (1986) Comparison of the
    Microtox test with the 96-hr LC50 test for the harpacticoid
     Nitocra  spinipes. Ecotoxicol Environ Saf, 11: 127-143.

    Tetra Tech Inc. (1986) Development of sediment quality values for
    puget sound. Bellevue, Washington, Tetra Tech Inc., vol 1,
    128 pp.

    Thurén A (1986) Determination of phthalates in aquatic
    environments. Bull Environ Contam Toxciol, 36(1): 33-40.

    Thurén A & Larsson P (1990) Phthalate esters in the Swedish
    atmosphere. Environ Sci Technol, 24(4):  554-559.

    Thurén A & Woin P (1991) Effects of phthalate esters on the
    locomotor activity of the freshwater amphipod  Gammarus pulex. 
    Bull Environ Contam Toxicol, 46(1): 159-166.

    Tomita I, Nakamura Y, & Yagi Y (1977) Phthalic acid esters in
    various foodstuffs and biological materials. Ecotoxicol Environ
    Saf, 1: 275-287.

    Tsuchiya K & Hattori K (1976) [Chromosomal study on human
    leucocyte cultures treated with phthalate acid ester.] Japan,
    Hokkaido Institute of Public Health (in Japanese).

    US EPA (1981) An exposure and risk assessment for phthalate
    esters. Washington, DC, US Environmental Protection Agency,
    Office of Water Regulations and Standards (EPA 440/4-81-020; NTIS
    Publication No. PB85-211936).

    US EPA (1982a) Aquatic fate process data for organic priority
    pollutants. Washington, DC, US Environmental Protection Agency
    (EPA 440/4-81-014; NTIS Publication No. PB87-169090).

    US EPA (1982b) Methods for organic chemical analysis of municipal
    and industrial wastewater. Cincinnati, Ohio, US Environmental
    Protection Agency, Environmental Monitoring and Support
    Laboratory (EPA 600/4-82-057; NTIS Publication No. PB83-201798).

    US EPA (1986a) Method 8060: Phthalate esters. In: Test methods
    for evaluating solid wastes, 3rd ed. Washington, DC, US
    Environmental Protection Agency, Office of Solid Waste and
    Emergency Response.

    US EPA (1986b) Method 8250: Gas chromatography/mass spectrometry
    for semivolatile organics - packed column technique. In: Test
    methods for evaluating solid wastes, 3rd ed. Washington, DC, US
    Environmental Protection Agency, Office of Solid Waste and
    Emergency Response.

    US EPA (1986c) Method 8270: Gas chromatography/mass spectrometry
    for semivolatile organics - capillary column technique. In: Test
    methods for evaluating solid wastes, 3rd ed. Washington, DC, US
    Environmental Protection Agency, Office of Solid Waste and
    Emergency Response.

    US EPA (1986d) Method 8410: Capillary column analysis of
    semivolatile organic compounds by gas chromatography/fourier
    transform infrared (GC/FT-IR) spectrometry. In: Test methods for
    evaluating solid wastes, 3rd ed. Washington, DC, US Environmental
    Protection Agency, Office of Solid Waste and Emergency Response.

    Virgin HI (1988) Accumulation of di-n-butylphthalate in plants
    and its effect on pigment and protein content. Physiol Plant,
    72(1): 190-196.

    Von Westernhagen H, Landolt M, Kocan R, Fürstenberg G, Janssen
    D, & Kremling K  (1987) Toxicity of sea-surface microlayer:
    effects on herring and turbot embryos. Mar Environ Res, 23:
    273-290.

    Waldock MJ (1983) Determination of phthalate esters in samples
    from the marine environment using gas chromatography mass
    spectrometry. Chem Ecol, 1: 261-277.

    Walker WW, Cripe CR, Pritchard PH, & Bourquin AW (1984) Dibutyl-
    phthalate degradation in estuarine and freshwater sites.
    Chemosphere, 13(12): 1283-1294.

    Walseth F & Nilsen OG (1984) Phthalate esters. II. Effects of
    inhaled dibutylphthalate on cytochrome P-450 mediated metabolism
    in rat liver and lung. Arch Toxicol, 55: 132-136.

    Walseth F & Nilsen OG (1986) Phthalate esters: Effects of orally
    administered dibutylphthalate on cytochrome P-450 mediated
    metabolism in rat liver and lung. Acta Pharmacol Toxicol, 59:
    263-269.

    Walseth F, Toftgard R, & Nilsen OG (1982) Phthalate esters I:
    Effects on cytochrome P-450 mediated metabolism in rat liver and
    lung, serum enzymatic activities and serum protein levels.
    Arch Toxicol, 50: 1-10.

    Wang JL, Liu P, & Qian Y (1995) Microbial degradation of di-n-
    butyl phthalate. Chemosphere, 31(9): 4051-4059.

    Ward TJ & Boeri RL (1991) Early life stage toxicity of di-n-butyl
    phthalate (DnBP) to the rainbow trout ( Oncorhynchus mykiss) 
    under flow-through conditions. Hampton, New Hampshire, Resource
    Analysts, Inc., EnviroSystems Division (Toxicity report submitted
    to the Chemical Manufacturers Association, Washington, DC).

    Webber MD & Lesage S (1989) Organic contaminants in Canadian
    municipal sludges. Waste Manage Res, 7: 63-82.

    Weschler C (1981) Identification of selected organics in the
    Arctic aerosol.  Atmos Environ, 15(8): 1365-1369.

    White RD, Carter DE, Earnest D, & Mueller J (1980)  bsorption and
    metabolism of three phthalate diesters by the rat small
    intestine. Food Cosmet Toxicol, 18: 383-386.

    White RD, Earnest DL, & Carter DE (1983) The effect of intestinal
    esterase inhibition on the  in vivo absorption and toxicity of
    di-n-butyl phthalate. Food Cosmet Toxicol, 21: 99-101.

    Williams DT (1973) Dibutyl- and di-(2-ethylhexyl)phthalate in
    fish. J Agric Food Chem, 21(6): 1128-1129.

    Williams DT & Blanchfield BJ (1975) The retention, distribution,
    excretion and metabolism of dibutyl phthalate-7-14C in the rat.
    J Agric Food Chem, 23(5): 854-858.

    Wilson WB, Giam CS, Goodwin TE, Aldrich A, Carpenter V, & Hrung
    YC (1978) The toxicity of phthalates to the marine dinoflagellate
     Gymnodinium breve. Bull Environ Contam Toxicol, 20(2):
    149-154.

    Wilson LG, Osborn MD, Olson KL, Maida SM, & Katz LT (1990) The
    ground water recharge and pollution potential of dry wells in
    Pima County, Arizona. Ground Water Monit Rev, 10: 114-121.

    Wine RN, Li LH, Barnes LH, Gulati OK, & Chapin RE (1977)
    Reproductive toxicity of di- n-butylphthalate in a continuous
    breeding protocol in SD rats. Environ Hlth Persp, 105(1):
    102-107.

    Wofford HW, Wilsey CD, Neff GS, Giam CS, & Neff JM (1981)
    Bioaccumulation and metabolism of phthalate esters by oysters,
    brown shrimp, and sheepshead minnows. Ecotoxicol Environ Saf,
    5: 202-210.

    Wolfe NL, Burns LA, & Steen WC (1980) Use of linear free energy
    relationships and an evaluative model to assess the fate and
    transport of phthalate esters in the aquatic environment.
    Chemosphere, 9: 393-402.

    Woodward KN (1988) Phthalate esters: Toxicity and metabolism,
    Volumes I and II. Boca Raton, Florida, CRC Press Inc.

    Woodward KN (1990) Phthalate esters, cystic kidney disease in
    animals and possible effects on human health: A review. Hum Exp
    Toxicol, 9: 397-401.

    Wu C, Pei X, Cao J, & Xue H (1993) A new pharmacological activity
    of dibutyl phthalate (DBP) on selective elimination of tumor
    cells from bone marrow. Leuk Res, 17(7): 557-560.

    Yagi Y, Tutikawa K, & Shimoi N (1976) Teratogenicity and
    mutagenicity of a phthalate ester. Teratology, 14: 259-260.

    Yagi Y, Nakamura Y, Tomita I, Tutikawa K, & Shimoi N (1978)
    Embryotoxicity of phthalate esters in mice. In: Plaa GL & Duncan
    WAM ed. Proceedings of the First International Congress on
    Toxicology. New York, London, San Francisco, Academic Press, pp
    590-591.

    Yamada A (1974) [Toxicity of phthalic acid esters and
    hepatotoxicity of di-(2-ehtyl hexyl) phthalate.] J Food Hyg Soc
    Jpn, 15: 147-152 (in Japanese, with English summary).

    Yan H, Ye C, & Yin C (1995) Kinetics of phthalate ester
    biodegradation by  Chlorella pyrenoidosa. Environ Toxicol Chem,
    14(6): 931-938.

    Yoshikawa K, Tanaka A, Yamada T, & Kurata H (1983) Mutagenicity
    study of nine monoalkyl phthalates and a dialkyl phthalate using
     Salmonella typhimurium and  Escherichia coli. Food Chem
    Toxicol, 21(2): 221-223.

    Yoshioka Y, Ose Y, & Sato T (1985) Testing for the toxicity of
    chemicals with  Tetrahymena pyriformis.  Sci Total Environ,
    43(1/2):  149-157.

    Yoshioka Y, Ose Y, & Sato T (1986) Correlation of the five test
    methods to assess chemical toxicity and relation to physical
    properties. Ecotoxicol Environ Saf, 12:15-21.

    Zeiger E, Haworth S, Speck W, & Mortelmans K (1982) Phthalate
    ester testing in the National Toxicology Program's environmental
    mutagenesis test development program. Environ Health Perspect,
    45: 99-101.

    Zeiger E, Haworth S, Mortelmans K, & Speck W (1985) Mutagenicity
    testing of di(2-ethylhexyl)phthalate and related chemicals in
     Salmonella. Environ Mutagen, 7: 213-232.

    Zhou Y, Fukuoka M, & Tanaka A (1990) Mechanisms of testicular
    atrophy induced by di- n-butyl phthalate in rats. Part 3.
    Changes in the activity of some enzymes in the Sertoli and germ
    cells, and in the levels of metal ions. J Appl Toxicol, 10(6):
    447-453.

    Ziogou K, Kirk PWW, & Lester JN (1989) Behaviour of phthalic acid
    esters during batch anaerobic digestion of sludge. Water Res,
    23(6): 743-748.

    Zitko V (1972) Determination, fate, and environmental levels of
    phthalate plasticizers. St. Andrews, N.B., Fisheries Research
    Board of Canada, Biological Station (Technical Report No. 344).

    Zurmühl T (1990) Development of a method for the determination of
    phthalate esters in sewage sludge including chromatographic
    separation from polychlorinated biphenyls, pesticides and
    polyaromatic hydrocarbons. Analyst, 115: 117.

    Zurmühl T, Durner W, & Herrmann R (1991) Transport of phthalate-
    esters in undisturbed and unsaturated soil columns. J Contam
    Hydrol, 8: 111-133.

    RESUME ET EVALUATION, CONCLUSIONS ET RECOMMANDATIONS

         Le phtalate de di- n-butyle (DBP) est un liquide inerte,
    incolore et de consistance huileuse, qui présente une faible
    tension de vapeur.  Soluble dans la plupart des solvants
    organiques, il n'est que légérement soluble dans l'eau. 
    L'analyse la plus sélective et la plus sensible des prélèvements
    effectués dans l'environnement  pour la recherche et le dosage du
    DBP et, plus généralement, des esters phtaliques, se fait par
    chromatographie en phase gazeuse avec détection par capture
    d'électrons ou spectrométrie de masse. Etant donné que des
    phtalates peuvent être présents sous la forme de plastifiants
    dans certaines pièces des instruments de mesure ou
    encore contaminer l'air du laboratoire, il faudra veiller tout
    particulièrement à éviter la contamination lors du prélèvement,
    de la conservation et de l'analyse des échantillons.

         Le DBP est principalement utilisé comme plastifiant de la
    nitrocellulose, de l'acétate et du chlorure de polyvinyle, comme
    lubrifiant des buses de bombes aérosol, comme agent antimoussant,
    comme adoucissant de la peau, comme plastifiant dans le vernis à
    ongles et les faux ongles ou encore dans les aérosols
    capillaires.  Le dosage du DBP présent dans l'atmosphère
    s'effectue en phase vapeur ou particulaire.  On pense que, pour
    une part non négligeable, le DBP est évacué de l'atmosphère par
    les précipitations ou par dépôt à sec.  Dans les eaux de surface,
    la majeure partie du  DBP est présente dans la phase liquide
    plutôt que dans les solides en suspension.  Il ne semble pas que
    le composé puisse s'évaporer du sol en quantités appréciables en
    raison de sa faible tension de vapeur et du fait qu'il est
    modérément adsorbé sur les particules du sol.

         La persistance du DBP dans l'air et les eaux de surface est
    relativement faible et sa demi-vie dans ces compartiments du
    milieu n'est que de quelques jours.  Il est rapidement biodégradé
    en aérobiose et beaucoup plus lentement en anaérobiose.  On a
    estimé que sa demi-vie dans le sol était du même ordre que dans
    l'air et l'eau, mais il semble, selon certaines études, que le
    DBP persiste plus longtemps dans le sol.  On peut s'attendre à
    une bioaccumulation importante du fait de la valeur élevée du
    coefficient de partage entre l'octanol et l'eau. Toutefois, sa
    métabolisation est rapide chez les poissons, aussi le facteur de
    bioconcentration a-t-il tendance à être plus faible que prévu. 
    La valeur la plus élevée (relativement au composé initial), à
    savoir 590, a été observée chez un cyprinidé d'Amérique du Nord,
     Pimephales promelas. Chez les animaux terrestres, il est peu
    probable qu'il y ait une bioamplification notable, à en juger
    d'après quelques données concernant les oiseaux et compte tenu du
    fait que la métabolisation et l'excrétion sont rapides chez les
    mammifères de laboratoire.

         Il n'est pas possible d'apprécier dans quelle mesure les
    anciennes données de surveillance peuvent être considérées comme
    fiables car, dans la littérature antérieure à 1980, on ne trouve
    que rarement mention des dispositions prises pour éviter la
    contamination des échantillons prélevés dans l'environnement. 
    Les données limitées dont on dispose au sujet des concentrations
    dans l'air ambiant indiquent que les valeurs moyennes sont
    généralement inférieures à 5 ng/m3.  Des études récentes ont
    montré que dans l'eau de pluie, la concentration moyenne allait
    de 0,2 à 1,4 µg/litre; des valeurs beaucoup plus faibles ont été
    mesurées dans des régions reculées.  Dans les eaux de surface, la
    concentration a tendance à être inférieure à 1 µg/litre;
    cependant on a relevé des valeurs beaucoup plus élevées dans des
    cours d'eau pollués (12 à 34 µg/litre). On ne possède que
    quelques données sur la concentration dans les eaux souterraines,
    les valeurs se situant entre 0,15 et 0,46 µg/litre.  Dans les
    effluents, la concentration du DBP peut aller jusqu'à 100
    µg/litre; elle varie de 0,2 à 200 µg/kg de poids sec dans les
    boues résiduaires.  Dans les sédiments, la concentration est en
    général inférieure à 1 mg/kg de poids sec; toutefois, dans les
    zones polluées, on a mesuré des concentrations pouvant aller
    jusqu'à 20 mg/kg. Selon les études portant sur la faune et la
    flore aquatiques, les concentrations auraient tendance à se
    situer à moins de 0,2 mg/kg de poids humide; néanmoins, des
    valeurs atteignant 35 mg/kg ont été relevées dans des secteurs
    pollués.

         Lors d'une enquête menée en Californie sur plus de 125
    résidences au cours de l'année 1990, on a relevé une
    concentration diurne moyenne de 420 ng/m3 dans l'air intérieur. 
    Quelques données canadiennes indiquent que la présence de DBP
    dans l'eau de boisson est plutôt rare et que sa concentration est
    inférieure à 1,0 µg/litre.  A Toronto, l'analyse d'un petit
    nombre d'échantillons a montré que la concentration du DBP y
    était de 14 ng/litre; on en a trouvé de 21 à 55 ng/litre dans des
    échantillons de plusieurs marques d'eau minérale en bouteille.

         Du DBP peut pénétrer dans les aliments par suite d'une
    contamination de l'environnement, mais la présence de ce composé
    dans une denrée alimentaire peut également être due à la
    migration du DBP de l'emballage vers son contenu.  Ce problème a
    été étudié à plusieurs reprises vers la fin des années 80.  Dans
    de nombreux pays, des précautions ont été prises pour réduire le
    passage, par lixiviation, des plastifiants de l'emballage dans le
    produit alimentaire. Ces mesures ont eu pour effet de réduire peu
    à peu  la teneur des aliments en DBP.  En 1986, on a effectué à
    Halifax une enquête sur le panier de la ménagère au cours de
    laquelle 98 produits alimentaires ont été étudiés.  La présence
    de DBP a été décelée dans du beurre (1,5 µg/g), du poisson d'eau
    douce (0,5 µg/g), des produits céréaliers (de non décelable à
    0,62 µg/g), des pommes de terre rôties (0,63 µg/g), de la salade

    de chou (0,11 µg/g), des bananes (0,12 µg/g), des airelles
    (0,09µg/g), des ananas (0,05 µg/g), de la margarine (0,64 µg/g),
    du sucre raffiné (0,2 µg/g), et des desserts à la gélatine
    (0,09 µg/g).

         Sur la base des données limitées dont on dispose, on peut
    dresser la liste suivante des principaux milieux par lesquels la
    population est exposée au DBP (par ordre d'importance
    décroissante et en fonction de la dose absorbée estimative):
    alimentation, air intérieur et eau de boisson. On estime que la
    dose absorbée quotidiennement à partir de l'alimentation et de
    l'air intérieur est respectivement égale à 7 µg/kg et 0,42 µg/kg
    de poids corporel.  Les doses absorbées journellement à partir de
    l'eau de boisson et de l'air ambiant sont très inférieures à ces
    valeurs, à savoir <0,02 µg/kg de poids corporel et 0,26 -
    0,36 ng/kg depoids corporel, respectivement.  En se basant sur
    ces valeurs, on peut calculer que la dose moyenne totale absorbée
    en une journée à partir de l'air, de l'eau de boisson et des
    aliments est égale à 7,4 µg/kg de poids corporel.  Toutefois, il
    est à noter que la dose absorbée à partir des aliments peu varier
    dans de larges proportions selon la nature et la quantité du
    produit emballé qui est consommé et également, selon les
    modalités d'utilisation de tel ou tel emballage au cours de la
    préparation du produit.  On estime qu'au Royaume-Uni, la dose
    maximale ainsi ingérée est probablement de 2 mg environ par
    personne et par jour (soit approximativement 31 µg/kg de poids
    corporel en une journée, pour une personne d'un poids moyen de
    64 kg).  Il y a également un risque d'exposition au DBP présent
    dans les cosmétiques, encore que dans ce cas, on ne dispose pas
    de données suffisantes pour évaluer la dose absorbée de cette
    manière.

         Les données provisoires les plus récentes fournies par
    l'enquête nationale sur l'exposition professionnelle qu'a menée
    le NIOSH indiquent qu'aux Etats-Unis, plus de 500 000
    travailleurs, dont 200 000 femmes, courent un risque d'exposition
    au DBP.  D'après les mesures effectuées sur un nombre limité de
    sites de ce pays, la concentration du DBP est généralement
    inférieure à la limite de détection (soit de 0,01 à 0,02 mg/m3),
    mais des valeurs plus élevées ont été relevées dans d'autres
    pays.

         Les études sur le rat montrent que le DBP est absorbé par la
    voie percutanée mais des travaux au cours desquels de la peau
    humaine a été exposée  in vitro ont révélé que cette dernière
    était moins perméable au DBP que la peau du rat.
    L'expérimentation animale (sur le rat) indique qu'une fois
    administré par voie orale ou intraveineuse, le DBP est rapidement
    résorbé au niveau des voies digestives et se répartit
    principalement dans le foie et les reins, avant d'être éliminé
    dans les urines sous forme de métabolites. Après inhalation, on
    le retrouve systématiquement à faible concentration dans
    l'encéphale.

         Les données disponibles montrent qu'après ingestion par des
    rats de laboratoire, le DBP est métabolisé par des esterases non
    spécifiques, principalement dans l'intestin grêle, pour donner du
    phtalate de mono- n-butyle (MBP) qui subit ensuite une oxydation
    limitée de sa chaîne latérale alkyle.  Le MBP est stable et le
    deuxième goupement ester résiste à l'hydrolyse.  Ce composé ainsi
    que les autres métabolites sont excrétés dans les urines sous
    forme de glucuro-conjugués.  On observe des différences
    interspécifiques relativement à l'excrétion des métabolites
    conjugués et non conjugués, notamment entre hamster et le rat,
    l'urine de ce dernier contenant davantage de MBP libre.  Aucune
    accumulation n'a été observée au niveau des divers organes.

         La palette des effets de l'exposition au DBP est analogue à
    celle que l'on observe avec les autres esters phtaliques, qui,
    chez les espèces sensibles, peuvent se réveler foetotoxiques et
    tératogènes et provoquer également une  hépatomégalie, un
    accroissement du nombre des peroxysomes hépatiques et des lésions
    testiculaires.

         Le DBP présente une faible toxicité aiguë pour le rat et la
    souris.  Après administration par voie orale à des rats on a
    obtenu, pour la DL50, des valeurs allant d'environ 8 g/kg  à au
    moins 20 g/kg de poids corporel.  Chez la souris, les valeurs
    vont d'environ 5 g/kg à environ 16 g/kg de poids corporel.  Chez
    le lapin, la DL50 cutanée est supérieure à 4 g/kg de poids
    corporel.  L'existence d'effets toxiques aigus consécutifs à
    l'inhalation de DBP n'a pu être documentée.  Chez les animaux de
    laboratoire, l'intoxication aiguë se manifeste par les signes
    suivants: réduction de l'activité, respiration difficile et perte
    de coordination.  Un travailleur qui avait été intoxiqué par
    suite de l'ingestion accidentelle de 10 g de DBP, s'est remis
    progressivement de son intoxication en l'espace de quinze jours,
    la récupération étant totale au bout d'un mois.

         Lors d'études toxicologiques au cours desquelles des rats
    ont reçu DBP à plusieurs reprises, on a observé, au bout d'une
    période de 5 à 21 jours, une prolifération des peroxysomes et une
    hépatomégalie aux doses supérieures ou égales à 420 mg/kg de
    poids corporel sur une journée.

         Lors d'études à plus long terme, les effets observés sur des
    rats ayant ingéré du DBP pendant des périodes allant jusqu'à 7
    mois consistaient notamment en une réduction du gain de poids à
    des doses quotidiennes supérieures ou égales à 250 mg/kg de poids
    corporel.  A des doses supérieures ou égales à 120 mg/kg de poids
    corporel, il y avait augmentation du poids relatif du foie. 
    Lorsque la dose quotidienne dépassait 279 mg/kg de poids
    corporel, on observait également une prolifération des
    peroxysomes et un accroisssement de l'activité des enzymes
    correspondantes.  Chez des rats Wistar qui avaient reçu des doses

    quotidiennes supérieures ou égales à 250 mg/kg de poids corporel,
    on a observé des signes de nécrose hépatique; en revanche ,
    aucune lésion de ce genre n'a été relevée chez des rats
    F-344 ou Sprague-Dawley soumis à des doses quotidiennes égales ou
    supérieures à 2 500 mg/kg de poids corporel. Aux doses
    quotidiennes de 250 et 571 mg/kg de poids corporel, on a observé
    chez le rat un certain nombre d'anomalies au niveau testiculaire,
    notamment des modifications affectant les enzymes et une
    dégénérescence des cellules germinales.  Les anomalies
    testiculaires observées varient considérablement d'une espèce à
    l'autre, les effets étant minimaux chez le hamster et la souris à
    des doses quotidiennes qui peuvent atteindre 2 000 mg/kg.  Une
    récente étude consistant en une exposition subchronique a permis
    de mettre en évidence, chez la souris, des effets sur le poids du
    corps et le poids des organes, de même que des modifications
    histologiques au niveau du foie, qui trahissent l'existence d'un
    stress métabolique; la dose sans effet observable pour ce type
    d'anomalie a été évaluée à 353 mg/kg de poids corporel.

         D'après les quelques données d'expérimentation animale dont
    on dispose, il semble que le DBP ne puisse guère provoquer
    d'irritation cutanée ou oculaire ni entraîner une
    sensibilisation. Chez l'homme, on connaît quelques cas de
    sensibilisation après exposition à du DBP, mais ces observations
    n'ont pas été confirmées par des études contrôlées sur un plus
    grand nombre d'individus.

         Dans le cadre d'un protocole d'élevage en continu, au cours
    duquel on a procédé à des croisements et à l'examen de la
    progéniture obtenue, des rats ont reçu une alimentation conentant
    0, 1000, 5000 ou 10 000 mg de DBP par kg de nourriture (soit
    l'équivalent quotidien de 0, 66, 320 et 651 mg de composé par kg
    de poids corporel). Dans la première génération, on a pu
    considérer comme un effet négatif sur le développement la
    réduction du poids corporel observée chez les ratons ayant reçu
    la dose médiane. On constatait également une réduction sensible
    du nombre de portées viables à toutes les doses. Dans la deuxième
    génération, les effets étaient plus graves, et consistaient en
    une réduction du poids des ratons dans tous les groupes, y
    compris celui qui avait reçu la dose la plus faible, en anomalies
    morphologiques (malformations du prépuce et du pénis,
    dégénérescence des tubes séminifères, et enfin, absence ou
    développement insuffisant de l'épididyme) dans les groupes soumis
    aux doses moyennes et fortes. Dans le groupe soumis à la dose la
    plus forte, on notait de graves effets sur la spermatogénèse,
    effets que l'on n'observait pas, en revanche, dans la génération
    parentale. Ces résultats donnent à penser que les effets nocifs
    du DBP sont plus marqués chez les animaux exposés au cours de
    leur phase de développement et de maturation que lorsqu'ils le
    sont uniquement à l'âge adulte.  Aucune valeur bien nette de la
    dose sans effet nocif observable (NOEL) n'a été tirée de cette

    étude. On estime en revanche que la dose la plus faible
    produisant un effet nocif (LOAEL) était égale à 66 mg/kg de poids
    corporel par jour.

         Les études dont on dispose montrent que le DBP est
    généralement foetotoxique, sans pour autant qu'il y ait atteinte
    de la mère.  Les données existantes indiquent également que ce
    composé est tératogène à forte dose, la sensibilité à cet effet
    dépendant du stade de développement et de la période
    d'administration.  Chez la souris, on constaté que le DBP
    provoquait une augmentation des résorptions et des morts foetales
    à partir de 400 mg/kg de poids corporel, cet effet étant lié à la
    dose.  Pour ces valeurs de la dose, on a également constaté chez
    la souris une réduction, liée à la dose, du poids foetal et du
    nombre de portées viables.

         On n'a pas effectué d'épreuves de cancérogénicité qui soient
    satisfaisantes. A la lumière des données disponibles, on peut
    penser que le DBP n'est pas génotoxique.

         Les produits chimiques qui provoquent la prolifération des
    peroxysomes constituent un groupe de substances souvent
    génératrices de cancers du foie, selon un mécanisme qui
    n'implique pas d'action toxique au niveau génique.  Leur mode
    d'action n'est pas encore élucidé, mais l'on sait cependant que
    l'apparition de la tumeur est précédée par une prolifération des
    peroxysomes et par une hépatomégalie.  Comme le DBP provoque la
    prolifération des peroxysomes, il n'est pas exclu qu'il puisse
    également provoquer des cancers du foie chez les rongeurs, encore
    qu'en ce qui concerne ces deux effets - prolifération des
    peroxysomes et hépatomégalie - il soit beaucoup moins actif que
    le DEHP.  Dans la mesure où il y a corrélation entre ces deux
    effets et le pouvoir cancérogène, on peut s'attendre à ce que le
    DBP soit un cancérogène beaucoup moins puissant que le DHEP et
    les méthodes actuelles de détermination biologique du pouvoir
    cancérogène ne permettraient probablement pas de mettre une telle
    activité en évidence.

         Les enquêtes épidémiologiques dont on a connaissance se
    limitent à l'étude du cas de travailleurs exposés à des mélanges
    de phtalates. Elles ne nous permettent pas de progresser dans
    l'élucidation des effets dus au seul DBP.

         On a vu que le DBP n'étant  pas génotoxique et ayant un
    pouvoir cancérogène moindre que celui du DEHP, les méthodes
    actuelles de mesure du pouvoir cancérogène ne révèleraient
    vraisemblablement aucune activité de ce type.  Il est donc peu
    probable qu'à la concentration où il se trouve dans
    l'environnement, ce composé contribue notablement à accroître le
    risque de cancer.

         C'est la voie alimentaire qui est, de loin, la principale
    voie d'exposition au DBP. D'ailleurs, les données toxicologiques
    relatives aux autres voies sont insuffisantes pour permettre une
    évaluation.  On a donc établi une valeur-guide pour la voie
    orale, même si l'objectif final doit être de ramener l'exposition
    totale de toutes origines à une valeur inférieure à la dose
    journalière tolérable.

         On n'a pas pu établir de valeur bien nette pour la dose sans
    effet nocif observable (NOAEL) pour les points d'aboutissement
    toxicologique jugés les plus appropriés à l'établissment de
    valeurs-guides (en l'occurence, les effets néfastes sur la
    reproduction et le développement). La dose la plus faible sans
    effet nocif observable (LOAEL) sur la reproduction et le
    développement a été fixée à 66 mg/kg de poids corporel par jour à
    la suite d'une étude au cours de laquelle les animaux étaient
    élevés en continu, avec cette réserve qu'à cette dose, les effets
    observés étaient modérés et probablement réversibles. En se
    basant sur ces données, on modérés et probablement réversibles.
    En se basant sur ces données, on a fixé à 66 µg/kg p.c. la dose
    journalière tolérable, compte tenu d'un facteur d'incertitude de
    1000 (un facteur 10 pour les variations interspécifiques, un
    facteur 10 pour les variations interindividuelles et un facteur
    10 pour l'extrapolation de la LOAEL à la NOAEL).

         Les renseignements dont on dispose sur l'écotoxicité du DBP
    comportent des données de toxicité aiguë et de toxicité chronique
    obtenues sur diverses espèces aquatiques à différents stades de
    la chaîne alimentaire.  Pour les algues d'eau douce, la valeur la
    plus faible de la CE50 à 96 h qui ait été obtenue est égale à
    750 µg de DBP par litre.  La valeur la plus faible de la CL50
    obtenue pour un invertébré aquatique (mysidé) est de 750 µg/litre
    et on a relevé une CE50 à 48 h de 760 µg/litre pour des larves
    de moucherons. Les études de toxicité chronique ont montré que
    l'espèce d'invertébré la plus sensible était  Daphnia magna, 
    avec une concentration sans effet observable à 21 jours (survie
    parentale) de 500 µg/litre.  Une épreuve non conventionnelle
    effectuée sur  Gammarus pulex a donné, pour la valeur de la
    concentration la plus faible produisant un effet observable à 10
    jours, le chiffre de 500 µg/litre et, pour la valeur de la
    concentration sans effet observable, le chiffre de 100 µg/litre,
    le critère retenu étant, dans les deux cas, la réduction de
    l'activité locomotrice.  Des épreuves de toxicité aiguë
    pratiquées sur des poissons ont permis de constater que la valeur
    la plus faible de la CL50 à 96 h était de 350 µg/litre pour une
    espèce dulçaquicole, la perche jaune  Perca flavescens, et de
    600 µg/litre pour un sparidé marin.  L'étude de toxicité
    chronique la plus sensible qui ait été pratiquée utilisait la
    truite arc-en-ciel et alle a donné une valeur de 100 µg/litre
    pour la concentration sans effet observable à 99 jours

    (croissance) et une valeur de 190 µg/litre pour la concentration
    la plus faible produisant un effet observable à 99 jours, le
    critère toxicologique retenu étant une réduction d'environ 27% de
    la croissance.

         La toxicité aiguë du DBP est faible pour les oiseaux.

         La concentration moyenne actuelle du DBP dans l'eau ne
    représente qu'un faible risque pour les organismes aquatiques. 
    Cependant, dans les cours d'eau très pollués, la marge de
    sécurité est beaucoup plus faible.  On ne dispose pas de données
    suffisantes pour évaluer le risque encouru par les organismes
    sédimenticoles. Compte tenu du niveau d'exposition actuel, le
    risque reste faible pour les oiseaux et les mammifères
    piscivores.

    RESUMEN Y EVALUACION, CONCLUSIONES Y RECOMENDACIONES

         El di- n-butil ftalato (DBF) es un líquido oleoso, incoloro
    e inerte, con una presión de vapor baja, soluble en la mayor
    parte de los disolventes orgánicos, pero sólo ligeramente en
    agua.  Las determinaciones analíticas más sensibles y selectivas
    de los ésteres del ácido ftálico en el medio ambiente, incluido
    el DBF, se logran mediante cromatografía de gases con detección
    por captura de electrones o espectrometría de masas.  Habida
    cuenta de que con frecuencia los ftalatos se encuentran como
    plastificantes en el equipo analítico y como contaminantes en el
    aire y los disolventes del laboratorio, hay que tener una gran
    precaución para evitar la contaminación durante la recogida, el
    almacenamiento y el análisis de las muestras.

         El DBF se utiliza principalmente como plastificante especial
    para la nitrocelulosa, el acetato de polivinilo y el cloruro de
    polivinilo, lubricante de válvulas de aerosoles, agente
    antiespumante, emoliente de la piel y plastificante de esmaltes y
    alargadores de uñas y pulverizadores para el pelo.

         Se ha determinado la concentración de DBF en la atmósfera,
    tanto en la fase de vapor como en la de partículas.  Se considera
    que el arrastre por la lluvia y la precipitación en seco ejercen
    una función importante en su eliminación de la atmósfera.  En las
    aguas superficiales, la mayor parte del DBF está presente en la
    fracción de agua más que en los sólidos suspendidos.  La
    volatilización a partir del suelo se supone insignificante,
    puesto que su presión de vapor es baja y la adsorción en el suelo
    moderada.

         El DBF es relativamente no persistente en el aire y las
    aguas superficiales y tiene una semivida en estos compartimentos
    de sólo unos días.  La biodegradación total es rápida en
    condiciones aerobias, pero mucho más lenta en anaerobiosis.  Para
    el suelo se ha pronosticado una semivida semejante a las del aire
    y el agua; sin embargo, algunos estudios indican que el DBF puede
    ser más persistente en el suelo.  Sería de esperar que el DBF se
    bioacumulara, debido a su elevado coeficiente de reparto
    octanol/agua.  No obstante, los peces lo metabolizan bastante
    fácilmente y, por consiguiente, los factores de bioconcentración
    tienden a ser más bajos de lo previsto.  El factor de
    bioconcentración máximo, basado en el compuesto precursor, es 590
    para  Pimephales promelas.  No es probable la bioamplificación
    en los animales terrestres, de acuerdo con los datos limitados en
    aves y con la rapidez del metabolismo y excreción que se ha
    observado en mamíferos de laboratorio.

         Raramente se han descrito medidas adoptadas para evitar la
    contaminación en los informes sobre las concentraciones de DBF en
    el medio ambiente publicados antes de 1980, por lo que no se
    puede evaluar la fiabilidad de los primeros datos de vigilancia.
    Hay datos limitados sobre concentraciones en el aire que indican

    que los niveles medios suelen ser inferiores a 5 ng/m3.  En
    estudios recientes se ha observado que las concentraciones medias
    en el agua de lluvia oscilaban entre 0,2 y 1,4 µg/litro; en zonas
    remotas se han detectado valores muchos más bajos. Las
    concentraciones medias en el agua superficial tienden a ser
    inferiores a 1 µg/litro; sin embargo, los niveles en ríos
    contaminados son mucho más elevados (12 a 34 µg/litro).  Se
    dispone de pocos datos sobre las concentraciones de DBF en el
    aguas freática, con valores medios de 0,15 a 0,46 µg/litro.  La
    concentración en efluentes alcanza hasta 100 µg/litro, mientras
    que en las aguas residuales varía de 0,2 a 200 mg/kg de peso
    seco.  Los niveles en los sedimentos son en general inferiores a
    1 mg/kg de peso seco; sin embargo, en zonas contaminadas se han
    medido concentraciones de hasta 20 mg/kg. En estudios realizados
    en la biota acuática se ha comprobado que las concentraciones
    medias de DBF tienden a ser menores de 0,2 mg/kg de peso seco;
    sin embargo, en zonas contaminadas se han medido concentraciones
    de hasta 35 mg/kg.

         En un estudio realizado en 125 hogares de California,
    Estados Unidos, en 1990, la concentración media durante el día en
    el aire de la casa era de 420 ng/m3.  Raramente se ha detectado
    DBF en el agua de bebida (<1,0 µg/litro), según datos limitados
    procedentes del Canadá. En un pequeño número de muestras de agua
    de bebida de Toronto, Canadá, la concentración media era
    14 ng/litro; las concentraciones en siete marcas de agua de
    manantial embotellada oscilaban entre 21 y 55 ng/litro.

         Además de su entrada mediante la contaminación del medio
    ambiente, el DBF puede estar presente en productos alimenticios
    como consecuencia de la migración desde el envase, aspecto que se
    investigó en varios estudios realizados a finales del decenio de
    1980.  En muchos países, se adoptaron precauciones para reducir
    la lixiviación de plastificantes de los envases y gracias a ello
    los niveles de DBF en los productos alimenticios han disminuido a
    lo largo del tiempo.  En un estudio sobre la cesta de la compra
    canadiense con 98 de muestras de tipos diferentes de alimentos
    realizado en Halifax en 1986, se detectó DBF en la mantequilla
    (1,5 µg/g), el pescado de agua dulce (0,5 µg/g), los productos a
    base de cereales (entre indetectable y 0,62 µg/g), las papas
    cocidas (0,63 µg/g), la ensalada de col (0,11 µg/g), los bananos
    (0,12 µg/g), los arándanos (0,09 µg/g), las piñas (0,05 µg/g), la
    margarina (0,64 µg/g), el azúcar blanco (0,2 µg/g) y el postre de
    gelatina (0,09 µg/g).

         Teniendo en cuenta los limitados datos disponibles, los
    medios de exposición principales al DBF para la población
    general, enumerados en orden de su importancia relativa según la
    ingestión estimada son los siguientes: alimentos, aire de
    espacios cerrados y agua de bebida.  La ingesta estimada en
    alimentos y en el aire de espacios cerrados es de 7 µg/kg de peso

    corporal al día y 0,42 µg/kg de peso corporal al día,
    respectivamente. Las ingestas con el agua de bebida y el aire del
    medio ambiente son considerablemente inferiores, <0,02 µg/kg de
    peso corporal al día y 0,26-0,36 ng/kg de peso corporal al día,
    respectivamente.  Habida cuenta de estas ingestas, se estima que
    la cantidad media total diaria ingerida con el aire, el agua de
    bebida y los alimentos es de 7,4 µg/kg de peso corporal al día.
    Hay que señalar, sin embargo, que la ingestión de DBF en la
    alimentación puede variar considerablemente, en función de la
    naturaleza y la cantidad de los alimentos envasados consumidos y
    el tipo de uso de los envoltorios de los alimentos en la
    preparación de la comida. Para el Reino Unido, la ingestión
    humana máxima probable de DBF de fuentes alimenticias se ha
    estimado en unos 2 mg/persona/día (aproximadamente 31 µg/kg de
    peso corporal/día, suponiendo un peso corporal medio de 64 kg). 
    Existe también la posibilidad de exposición al DBF a través de
    los cosméticos, aunque los datos disponibles son insuficientes
    para cuantificar la ingestión a partir de esta fuente.

         Los datos provisionales más recientes de la Encuesta
    Nacional de Exposición Profesional NIOSH señalan que en los
    Estados Unidos hay más de 500 000 trabajadores, incluidas 200 000
    mujeres, potencialmente expuestos al DBF.  Teniendo en cuenta las
    determinaciones en un número limitado de puestos de trabajo en
    los Estados Unidos, las concentraciones son en general inferiores
    al límite de detección (es decir, 0,01-0,02 mg/m3), si bien se
    ha informado de niveles más elevados en algunos países.

         En estudios con ratas, se ha observado que el DBF se absorbe
    a través de la piel, aunque en estudios  in vitro la piel humana
    ha resultado menos permeable que la de rata a este compuesto.  En
    estudios con animales de laboratorio se ha advertido que, tras la
    administración oral o intravenosa, el DBF se absorbe rápidamente
    del tracto gastrointestinal, se distribuye fundamentalmente en el
    hígado y los riñones y se excreta en la orina como metabolitos. 
    Tras la inhalación se detectaron constantemente concentraciones
    bajas en el cerebro.

         Los datos disponibles indican que en ratas, tras la
    ingestión, el DBF se metaboliza mediante la acción de esterasas
    inespecíficas, sobre todo en el intestino delgado, para producir
    mono- n-butil ftalato (MBP), con la posterior oxidación
    bioquímica limitada de la cadena alcalina lateral del MBP.  Este
    compuesto es estable y resistente a la hidrólisis del segundo
    grupo éster.  El MBP y otros metabolitos se excretan en la orina,
    principalmente como conjugados de glucurónidos.  Se han observado
    especies diferentes en la excreción de metabolitos conjugados y
    no conjugados del DBF en la orina de rata y hámster, con más MBP
    libre en la rata que en el hámster. No se ha observado
    acumulación en ningún órgano.

         El perfil de los efectos tras la exposición al DBF es
    semejante al de otros ésteres de ftalatos, que en especies
    susceptibles puede inducir hepatomegalia, aumento del número de
    peroxisomas hepáticos, fetotoxicidad, teratogenicidad y daños
    testiculares.

         La toxicidad aguda del DBF en ratas y ratones es baja.  Los
    valores de la DL50 notificados tras la administración oral a
    ratas oscilan entre alrededor de 8 g/kg de peso corporal y por lo
    menos 20 g/kg de peso corporal; en ratones, los valores son
    aproximadamente de 5 g/kg de peso corporal y 16 g/kg de peso
    corporal.  La DL50 por vía cutánea en conejos es >4 g/kg de
    peso corporal.  No hay datos de toxicidad aguda tras la
    inhalación de DBF.  Los signos de toxicidad aguda observados en
    los animales de laboratorio incluyen depresión de la actividad,
    respiración fatigosa y falta de coordinación.  En un caso de
    intoxicación accidental de un trabajador que ingirió alrededor de
    10 g de DBF, la recuperación fue gradual en un plazo de dos
    semanas y completa al cabo de un mes.

         En estudios de toxicidad de corta duración con dosis
    repetidas, los efectos en ratas tras la administración oral
    durante un período de 5 a 21 días con los niveles más bajos
    fueron proliferación de peroxisomas y hepatomegalia a dosis de
    420 mg/kg de peso corporal al día o más.

         En estudios más prolongados, los efectos observados en ratas
    tras la ingestión de DBF durante períodos de hasta siete meses
    fueron un aumento reducido de peso a dosis de 250 mg/kg de peso
    corporal al día o más.  Se ha observado un aumento en el peso
    relativo del hígado a dosis de 120 mg/kg de peso corporal o más. 
    A dosis de 279 mg/kg de peso corporal o más se ha registrado
    proliferación de peroxisomas, con aumento de su actividad
    enzimática.  Se ha informado de cambios hepáticos necróticos en
    ratas Wistar a dosis de 250 mg/kg de peso corporal por día o más,
    pero no en ratas F-344 o Sprague-Dawley expuestas a dosis de
    hasta 2500 mg/kg de peso corporal al día.  Se han advertido
    alteraciones en las enzimas testiculares y degeneración de las
    células germinales testiculares de ratas con dosis de 250 y
    571 mg/kg de peso corporal al día.  Los efectos en los testículos
    tras la exposición al DBF son muy diferentes en las distintas
    especies, habiéndose observado efectos mínimos en ratones y
    hámster a dosis de hasta 2000 mg/kg de peso corporal al día.  En
    un bioensayo subcrónico reciente realizado en ratones, se han
    descrito efectos en el peso del cuerpo y los órganos y
    alteraciones histológicas del hígado, factor indicativo de
    tensión metabólica, por lo cual el NOEL fue 353 mg/kg de peso
    corporal al día.

         Teniendo en cuenta los limitados datos disponibles en
    especies animales, el DBF parece tener escaso potencial para
    irritar la piel o los ojos o inducir sensibilización.  En el ser
    humano se ha informado de un pequeño número de casos de
    sensibilización tras la exposición al DBF, aunque esto no se vio
    confirmado en estudios controlados realizados con un número más
    elevado de personas, notificados solamente en resultados
    secundarios.

         En un protocolo continuo de reproducción, que comprendió
    fases de apareamiento cruzado y de evaluación de la descendencia,
    se expusieron ratas a 0, 1000, 5000 o 10 000 mg de DBF en la
    alimentación (dosis equivalentes a 0, 66, 320 y 651 mg/kg de peso
    corporal por día).  En la primera generación, la reducción del
    peso de las crías en el grupo expuesto a la dosis intermedia, en
    ausencia de cualquier efecto adverso en el peso materno, pudo
    considerarse como efecto de toxicidad sobre el desarrollo.  Hubo
    también una reducción significativa del número de crías vivas en
    cada camada para los tres niveles de dosis.  En la segunda
    generación los efectos fueron más serios: reducción del peso de
    las crías en todos los grupos, incluido el grupo expuesto a la
    dosis baja; defectos estructurales (tales como malformaciones
    prepuciales/peneanas, degeneración de los túbulos seminíferos y
    ausencia o subdesarrollo de los epidídimos) en los grupos
    sometidos a la dosis intermedia/ alta; y efectos graves sobre la
    espermatogénesis en el grupo expuesto a la dosis alta, que no se
    observaron en los progenitores.  Estos resultados dan a entender
    que los efectos adversos del DBF son más acusados en los animales
    expuestos durante las fases de desarrollo y maduración que en los
    expuestos solo en la edad adulta.  No se estableció ningún NOEL
    en ese estudio. El nivel inferior con efectos adversos observados
    (LOAEL) se consideró que fue 66 mg/kg de peso corporal por día.

         En los estudios disponibles se ha puesto de manifiesto que
    el DBF suele inducir efectos fetotóxicos en ausencia de toxicidad
    materna.  Los datos existentes indican asimismo que el DBF es
    teratogénico a dosis elevadas y que la susceptibilidad a la
    teratogénesis varía en función de la fase de desarrollo y del
    período de administración.  La administración oral de DBF a
    ratones en dosis de 400 mg/kg de peso corporal o superiores
    produjo un aumento dependiente de la dosis en el número de
    reabsorciones y de muertes fetales.  Con estas mismas dosis se
    observó también en ratones una disminución dependiente de la
    dosis del peso fetal y del número de crías viable.

         No se han realizado bioensayos adecuados de carcinogénesis
    para el DBF.  Las pruebas disponibles indican que el DBF no es
    genotóxico.

         En general, los productos químicos que causan proliferación
    de los peroxisomas son con frecuencia hepatocarcinógenos mediante
    una acción no genotóxica.  Si bien no se conoce todavía el
    mecanismo de acción, la formación de tumores va precedida de
    proliferación de peroxisomas y hepatomegalia.  Habida cuenta de
    que el DBF produce proliferación de peroxisomas, es posible que
    pudiera ser un carcinógeno hepático en roedores, aunque como
    inductor de hepatomegalia y proliferación de peroxisomas es mucho
    más débil que el DEHF.  En la medida en que existe correlación
    entre la hepatomegalia y la formación de peroxisomas por una
    parte y la capacidad carcinógena por otra, cabe prever que el DBF
    será un carcinógeno menos potente que el DEHF y probablemente no
    mostrará actividad si la medición se realiza con las metodologías
    actuales de bioensayo del cáncer.

         Las investigaciones epidemiológicas identificadas se limitan
    a las de trabajadores expuestos a mezclas de ftalatos.  Estos
    estudios no contribuyen a mejorar nuestros conocimientos sobre
    los efectos asociados al DBF aislado.

         Puesto que el DBF no es genotóxico y se supone que será
    menos carcinógeno que el DEHF, probablemente no mostraría
    actividad si la medición se realizara utilizando las metodologías
    actuales de bioensayo del cáncer.  Así pues, no es probable que
    el DBF presente un aumento significativo del riesgo de cáncer en
    las concentraciones a las que habitualmente se encuentra en el
    medio ambiente.

         La ingestión es con diferencia la vía principal de
    exposición al DBF; además, los datos toxicológicos de las demás
    vías de administración son insuficientes para su evaluación.  Por
    consiguiente, se ha preparado un valor orientativo para la vía
    oral, aunque el objetivo último debería ser la reducción de la
    exposición total a todas las fuentes para lograr una ingesta
    diaria inferior a la tolerable.

         No se estableció ningún claro nivel sin efectos adversos
    observados (NOAEL) para los puntos finales considerados los más
    adecuados para obtener los valores de orientación (es decir, la
    toxicidad en el desarrollo y la reproducción). Se consideró que
    el LOAEL resultante de un estudio continuo de reproducción para
    la toxicidad en el desarrollo y la reproducción fue 66 mg/kg de
    peso corporal al día, aunque los efectos observados a ese nivel
    de dosis fueron moderados y probablemente reversibles.  A partir
    de esos datos, se ha obtenido una ingesta diaria tolerable de
    66 µg/kg de peso corporal al día, incorporando un factor de
    incertidumbre de 1000 (× 10 para la variación entre especies, ×
    10 para la variación entre individuos, y × 10 para la
    extrapolación del LOAEL al NOAEL).

         La información sobre la ecotoxicidad del DBF consiste en
    datos sobre toxicidad aguda y crónica para varias especies de
    distintos niveles tróficos del medio ambiente acuático.  La CE50
    más baja descrita para algas de agua dulce a las 96 horas fue de
    750 µg de DBF/litro.  Los valores más pequeños obtenidos en las
    pruebas de toxicidad aguda en invertebrados acuáticos fueron una
    DL50 a las 96 horas de 750 µg/litro (mísidos) y una CE50 a las
    48 horas de 760 µg/litro (larvas de mosca enana).  En estudios de
    toxicidad crónica, la especie más sensible de invertebrado fue
     Daphnia magna, con una NOEC (supervivencia de los padres) a
    los 21 días de 500 µg/litro.  En una prueba no normalizada con un
    antípodo ( Gammarus pulex) se obtuvo una LOEC a los 10 días de
    500 µg/litro y una NOEC de 100 µg/litro, basados ambos en una
    reducción de la actividad locomotriz. En pruebas de toxicidad
    aguda con peces, la CL50 más baja a las 96 horas notificada para
    una especie de agua dulce fue de 350 µg/litro (perca canadiense)
    y para una especie marina de 600 µg/litro (sargo chopa).  El
    estudio de toxicidad crónica más sensible se basó en la trucha
    irisada, con una NOEC (crecimiento) a los 99 días de 100 µg/litro
    y una LOEC a los 99 días de 190 µg/litro (reducción del
    crecimiento de alrededor del 27 por ciento).

         La toxicidad aguda del DBF para las aves es baja.

         El riesgo para los organismos acuáticos asociado a las
    concentraciones medias presentes en las aguas superficiales es
    bajo. Sin embargo, el margen de inocuidad en ríos muy
    contaminados es mucho más pequeño.  No se dispone de datos
    adecuados que permitan evaluar el riesgo del DBF para los
    organismos que viven en sedimentos.  Se puede concluir que, con
    los niveles actuales, el riesgo para las aves y los mamíferos que
    se alimentan de peces es bajo.
    


See Also:
        Dibutyl phthalate (CHEMINFO)
        Dibutyl phthalate (ICSC)