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 The International Programme on Chemical Safety (IPCS) is a joint venture of the United Nations Environment Programme, the International Labour Organisation, and the World Health Organization. The main objective of the IPCS is to carry out and disseminate evaluations of the effects of chemicals on human health and the quality of the environment. Supporting activities include the development of epidemiological, experimental laboratory, and risk-assessment methods that could produce internationally comparable results, and the development of manpower in the field of toxicology. Other activities carried out by the IPCS include the development of know-how for coping with chemical accidents, coordination of laboratory testing and epidemiological studies, and promotion of research on the mechanisms of the biological action of chemicals. WHO Library Cataloguing in Publication Data 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 The World Health Organization welcomes requests for permission to reproduce or translate its publications, in part or in full. Applications and enquiries should be addressed to the Office of Publications, World Health Organization, Geneva, Switzerland, which will be glad to provide the latest information on any changes made to the text, plans for new editions, and reprints and translations already available. (c) World Health Organization 1997 Publications of the World Health Organization enjoy copyright protection in accordance with the provisions of Protocol 2 of the Universal Copyright Convention. All rights reserved. The designations employed and the presentation of the material in this publication do not imply the expression of any opinion whatsoever on the part of the Secretariat of the World Health Organization concerning the legal status of any country, territory, city or area or of its authorities, or concerning the delimitation of its frontiers or boundaries. The mention of specific companies or of certain manufacturers' products does not imply that they are endorsed or recommended by the World Health Organization in preference to others of a similar nature that are not mentioned. Errors and omissions excepted, the names of proprietary products are distinguished by initial capital letters. 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 22.214.171.124 Surface water 126.96.36.199 Groundwater 188.8.131.52 Seawater 184.108.40.206 Precipitation 220.127.116.11 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 18.104.22.168 In vivo studies 22.214.171.124 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 126.96.36.199 Testicular effects 188.8.131.52 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 184.108.40.206 Algae 220.127.116.11 Invertebrates 18.104.22.168 Vertebrates 9.1.3. Terrestrial organisms 22.214.171.124 Plants 126.96.36.199 Invertebrates 188.8.131.52 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 Every effort has been made to present information in the criteria monographs as accurately as possible without unduly delaying their publication. In the interest of all users of the Environmental Health Criteria monographs, readers are requested to communicate any errors that may have occurred to the Director of the International Programme on Chemical Safety, World Health Organization, Geneva, Switzerland, in order that they may be included in corrigenda. * * * A detailed data profile and a legal file can be obtained from the International Register of Potentially Toxic Chemicals, Case postale 356, 1219 Châtelaine, Geneva, Switzerland (Telephone No. 9799111). * * * This publication was made possible by grant number 5 U01 ES02617- 15 from the National Institute of Environmental Health Sciences, National Institutes of Health, USA, and by financial support from the European Commission. Environmental Health Criteria PREAMBLE Objectives In 1973 the WHO Environmental Health Criteria Programme was initiated with the following objectives: (i) to assess information on the relationship between exposure to environmental pollutants and human health, and to provide guidelines for setting exposure limits; (ii) to identify new or potential pollutants; (iii) to identify gaps in knowledge concerning the health effects of pollutants; (iv) to promote the harmonization of toxicological and epidemiological methods in order to have internationally comparable results. The first Environmental Health Criteria (EHC) monograph, on mercury, was published in 1976 and since that time an ever-increasing number of assessments of chemicals and of physical effects have been produced. In addition, many EHC monographs have been devoted to evaluating toxicological methodology, e.g., for genetic, neurotoxic, teratogenic and nephrotoxic effects. 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It is accepted that the following criteria should initiate the updating of an EHC monograph: new data are available that would substantially change the evaluation; there is public concern for health or environmental effects of the agent because of greater exposure; an appreciable time period has elapsed since the last evaluation. All Participating Institutions are informed, through the EHC progress report, of the authors and institutions proposed for the drafting of the documents. A comprehensive file of all comments received on drafts of each EHC monograph is maintained and is available on request. The Chairpersons of Task Groups are briefed before each meeting on their role and responsibility in ensuring that these rules are followed. 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 g/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 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 184.108.40.206 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 g/litre) and Etherow (32.5 and 23.5 g/litre) in Manchester, United Kingdom (Fatoki & Vernon, 1990). Both rivers received discharges from factories making plastic products. 220.127.116.11 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. 18.104.22.168 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). 22.214.171.124 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. 126.96.36.199 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 g 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 g/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 g/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 188.8.131.52 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). 184.108.40.206 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 g/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 g/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 220.127.116.11 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, -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). 18.104.22.168 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 g/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 22.214.171.124 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). 126.96.36.199 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. 188.8.131.52 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 184.108.40.206 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). 220.127.116.11 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. 18.104.22.168 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 g/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 g/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) 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. 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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)