Di-n-butyl phthalate (EHC 189, 1997)

UNITED NATIONS ENVIRONMENT PROGRAMME
INTERNATIONAL LABOUR ORGANISATION
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

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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

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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.

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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 K_ow) 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

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 K_oc 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 samples^a

Sample matrix Sample preparation Analytical Sample detection
methods^b 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
methods^b 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/litre^d No data US EPA (1986d)

Wastes, Extract from sample, cleanup HRGC/FTIR 10 µg/litre^d 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
method^a 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 BCF^a Reference
concentration (days)
(µg/litre)

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

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

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

Scud 0.10 14 1400^c 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 BCF^a 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 7 720^c Mayer & Sanders (1973)
(Chironomus
plumosus)

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

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

Fathead minnow 34.8 11 590^d 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
Canada^a. Based upon the average daily consumption of various
foodstuffs by adults^b, 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