Review of Environmental Fate and Effects of Selected
Phthalate Esters

6 Di-n-butyl Phthalate (DBP)

6.1 Physico-chemical properties
6.1.1 Water solubility
6.1.2 Octanol-water partition coefficient
6.1.3 Summary
6.2 Environmental concentrations and fate
6.2.1 Concentrations in the environment
6.2.2 Abiotic degradation
6.2.3 Biodegradation
6.2.4 Bioaccumulation
6.2.5 Summary and conclusion
6.3 Effects
6.3.1 Terrestrial organisms
6.3.2 Toxicity to micro-organisms
6.3.3 Toxicity to algae
6.3.4 Toxicity to invertebrates
6.3.5 Toxicity to fish
6.3.6 Estrogenic effects
6.3.7 Summary and conclusions
6.4 Environmental hazard classification
6.5 PNEC for the aquatic compartment
 

The largest usage of DBP in general is as a plasticizer in resins and polymers such as polyvinyl chloride. Furthermore, DBP is used in other consumer products such as cosmetics: a perfume solvent and fixative, a suspension agent for solids in aerosols, a lubricant for aerosol valves, an antifoamer, a skin emollient and a plasticizer in nail polish, fingernail elongators and hair spray /63/.

6.1 Physico-chemical properties

DBP (C₁₆H₂₂O₄), CAS No.: 84-74-2, with a alkyl chain length of 4,4 /1/ is a colourless oily liquid. The molecular weight is 278.4 g/mol. DBP has a melting point of about -35°C and a boiling point at 340°C /6/. The density is 1.042 g/ml and the vapour pressure is 2.7·10-5 mmHg at 25°C /1/.

6.1.1 Water solubility

DBP is a low molecular weight phthalate and experimental determinations of water solubilities are believed to be reliable for lower molecular weight phthalates. For DBP, several aqueous solubility data are referred to in the literature ranging from 3.25 to 13.0 mg/l. The water solubility has been calculated to 1.53 mg/l /7/. In a literature review by Staples et al. /1/, it was concluded that a water solubility of about 11.2 mg/l was the most likely value based on available evidence.

6.1.2 Octanol-water partition coefficient

As for solubility, there are several different values in the literature for the octanol-water partition coefficient K_ow, differing by a factor of 25. Reported log K_ow values for DBP range from 3.74 to 5.15. Compared to these values, a log K_ow of 4.72 has been calculated /7/. According to Staples et al. /1/, a log K_ow of 4.45 is the most likely value based on available evidence.

6.1.3 Summary

The approximate physico-chemical properties on DBP are summarized in Table 6.1.

Table 6.1
Physico-chemical properties of Dibutyl Phthalate (DBP)

6.2 Environmental concentrations and fate

6.2.1 Concentrations in the environment

The content of DBP in wastewater and sewage sludge from Danish treatment plants has been measured at several occasions during recent years. An overview of the results is given in Table 6.2.

Table 6.2
link to table

It is a general picture that a high removal of DBP from the wastewater is found during the waste water treatment. Mass balances show that from 1 to 40% of the amount in the inlet water is found in sludge /9/, and it is indicated that the amount of DBP is reduced by anaerobic degradation.

A large number of data on concentrations of DBP in the environment is reported in TNO & RIVM (1997) /63/ and Györkös (1996) /3/ and only a brief overview will be given here. The quality of the analyses has not been evaluated in the present report, but it is expected that the data have been validated by TNO & RIVM (1997) /63/. More detailed information can be found in TNO & RIVM (1997) /63/ and references cited therein.

DBP has been detected in rivers, estuaries and sea water. In sea water, DBP concentrations in the range from 0.046 to 3.4 µg/l have been determined, while concentrations in estuaries in the range from 0.011 to 4.8 µg/l have been determined. The highest concentrations of DPB have been measured in rivers - 0.001 to 622.9 µg/l - thus the variation is considerable. Furthermore, DBP has been detected in ground water at concentrations ranging from 0.2 to 2,249 µg/l. The highest values have been characterised at a former waste disposal lagoon related to a chemical company, which had manufactured more than 200 different chemicals for 50 years /3/.

In 12 different sediment samples, the mean measured concentrations of DBP range from 0.001 to 2.2 mg/kg (dry weight basis) /63/. A calculated regional sediment concentration of 0.5 mg/kg (wet weight basis) equals the upper limit of the measured range and has been used for the risk characterization at a regional scale in TNO & RIVM (1997) /63/. In Györkös (1996) /3/, however, concentrations of up to 100 mg/kg have been described.

A very limited and not representative set of monitoring data on soil, ranging from <0.1 to 0.175 mg/kg, is available. Thus in TNO & RIVM (1997) /63/, a calculated soil concentration of 0.02 mg/kg has been used for the risk characterization of the terrestrial compartment.

Very limited data are available but indicate that the regional DBP concentrations in the EU range from 0.00023 to 0.056 µg/m³. The calculated regional PEC of 0.007 µg/m³ was found to be of the same order of magnitude and thus used for the risk characterisation in TNO & RIVM (1997) /63/.

DBP has been detected in several marine fish species. Concentrations of DBP in fish livers have been measured in the range from below the detection limit to 11700 µg/kg dry weight. In fish muscle tissue, DBP has been detected in the range from below the detection limit to 530 µg/kg dry weight. In aquatic invertebrates, terrestrial invertebrates, terrestrial plants and aquatic plants DBP has been measured in concentrations up to 500 µg/kg dry weight, 1750 µg/kg dry weight, 1557 µg/kg dry weight and 1900 µg/kg dry weight, respectively.

6.2.2 Abiotic degradation

Wolfe et al. (1980) /2/ measured the hydrolysis rate constant of DBP and estimated a half-life of 22 years at alkaline conditions. The hydrolysis half-life at neutral pH and 25°C is estimated to 3.4 years /7/.

No experimental data on photodegradation of DBP are available. Estimated photodegradation half-lives in the atmosphere are in the range from 0.6 to 6 days /4, 7/. In the aquatic environment, only insignificant photodegradation is expected /1, 3/.

6.2.3 Biodegradation

Scholz et al. (1997) /65/ investigated the ready biodegradability of DBP in the Modified Sturm test (OECD 301B) and found a degradation of 81% after incubation for 28 days. Györkös (1996) /3/ refers to a value of BOD₅ = 63% determined with non-adapted micro-organisms.

Sugatt et al. (1984) /14/ using acclimated inocula demonstrated a biodegradability of DBP of 57% after 28 days.

Staples et al. (1996) /1/ have reviewed the biodegradability of DBP and referred to numerous studies showing a primary biodegradability between 50% and 100%.

Howard (1998) /11/ refers to an investigation showing 60-70% removal in three sewage treatment plants using activated sludge. This complies with the measured removals in Danish treatment plants (Table 6.2).

Staples et al. (1996) /1/ refer to tests on anaerobic biodegradability of DBP showing a primary biodegradability of 66% to 100% and an ultimate biodegradability of 0% to 100% at 22-37°C and incubation for up to 140 days. Györkös (1996) /3/ refers to a study by Battersby & Wilson (1989) /66/ showing 24% mineralization of DBP after 77 days in diluted anaerobic sewage sludge. Furthermore, Györkös (1996) /3/ refers to a study by Shelton et al. (1984) /67/ demonstrating more than 90% degradation of DBP in undiluted sludge within 7 days while 40 days were needed in order to attain this level in diluted sludge containing 10% inoculum.

Johnson et al. (1984) /68/ investigated the biodegradation of DBP in sediment that was pre-exposed for 28 days before the start of the experiment. After 14 days of incubation under aerobic conditions, a primary degradation of 85% was determined. Further experiments demonstrated that the same degree of degradation (70-73% after 14 days) was reached at various concentrations from 0.08 to 8 mg/l. Finally, a significantly longer lag-phase was observed at low temperatures (5°C and 12°C) compared with higher temperatures (22°C and 28°C), and increasing degradation was observed at increasing temperatures from about 50% to 100%.

6.2.4 Bioaccumulation

In the review given by /1/, several bioaccumulation studies with algae, crustaceans, fish and insects are reported. A great variability exists between BCFs reported on total ¹⁴C and parent phthalate ester, respectively. Below, a number of studies on bioaccumulation of DBP referred to in Staples et al. (1996) /1/ are given.

Casserly et al. (1983) /69/ determined the BCF after 1 day of static exposure based on the parent compound for freshwater algae (Selenastrum capricornutum), BCF: 5475 and 1324, respectively.

For crustaceans, only data on total ¹⁴C exist. For Palaemontes kadiakensis, a BCF of 750 (3 days of exposure) was found, however, steady state was not reached in this study /70/. For scud, BCFs of 1485 /71/ and 185 (10 days of exposure) /72/, respectively, have been found. Finally, for artemia, a BCF of 345 has been reported by Hudson et al. (1981) /73/ (exposure period not known).

Call et al. (1993) /74/ found a total BCF for Fathead minnow of 2068 (11 days of exposure with an exposure concentration of 5 µg/l). Calculating from the fraction of total radioactivity reported as parent compound after 11 days, the BCF of the parent compound was determined to 167.

Mayer & Sanders (1973) /71/ reported a BCF based on total BCF for Hexagenia bilineata of 714 (exposure concentration: 0.08 µg/l and with a static renewal test procedure) and for Chironomus plumosus, a total BCF of 700 was found (exposure concentration: 0.18 µg/l and a static renewal test procedure). Sanders et al. (1973) /70/ reported a total BCF of 458 for Ischnura verticalis (exposure concentration: 0.1 µg/l and a static renewal test procedure).

6.2.5 Summary and conclusion

DBP seems to be efficiently removed from waste waters based on the low outlet concentrations compared to inlet concentrations in sewage treatment plants. However, large differences exist between different treatment plants. Degradation may account for some of the reduction seen but accumulation in sludge may be important based on the high concentrations of DBP measured in some sludge samples.

DBP has been detected in soil water (rivers, estuaries and seawater) and ground water at variable concentrations. In aquatic environments, DBP concentrations of up to 623 µg/l have been measured while concentrations as high as up to 2,249 µg/l have been detected in ground water. In sediments, DBP has been found in concentrations of up to 100,000 µg/kg. In biota, DBP has been measured in concentrations of up to 11,700 µg/kg, 500 µg/kg, 1,750 µg/kg, 1,557 µg/kg and 1,900 µg/kg dry matter, in fish, aquatic invertebrates, terrestrial invertebrates, terrestrial plants and aquatic plants, respectively.

Hydrolysis and photodegradation are not significant degradation routes of DBP in the aquatic environment.

DBP is readily biodegradable in standard laboratory tests and, hence, the substance is expected to be mineralized rapidly in the aerobic part of a sewage treatment plant. The studies of anaerobic biodegradability indicate a potential for mineralization of DBP during anaerobic treatment of sludge.

DBP is bioaccumulating in aquatic biota, which is demonstrated by the experimentally derived BCF values of up to 2,125 for fish.

6.3 Effects

6.3.1 Terrestrial organisms

Callahan et al. (1994) /75/ reported an LC₅₀ (48 h) value for DBP of 74 µg/kg^-1 when tested on the earthworm, Eisenia fetida, by use of a standard contact test.

6.3.2 Toxicity to micro-organisms

The toxicity studies with micro-organisms are summarized in Table 6.3. The table contains data on both bacteria and protozoa.

Table 6.3
link to table

In the Pseudomonas putida test the effect of DBP was found at concentrations above the water solubility of the substance.

6.3.3 Toxicity to algae

The toxicity studies with DBP for freshwater and marine algae are summarized in Table 6.4.

Table 6.4
link to table

Technical shortcomings have been described for the Gymnodinium test. Furthermore, from the results obtained, it can be seen that the reproducibility of the test is very poor (EC₅₀: 0.0032-0.2 mg/l). For this reason the test will not be used in the further evaluation of DBP.

6.3.4 Toxicity to invertebrates

The short-term toxicity data on DBP to freshwater and marine invertebrates are presented in Table 6.5. and the long-term toxicity data on DBP to freshwater and marine invertebrates are presented in Table 6.6.

Table 6.5
link to table

Table 6.6
link to table

6.3.5 Toxicity to fish

The short-term toxicity data on DBP to freshwater and marine fish are presented in Table 6.7 and the long-term toxicity data on DBP to freshwater and marine fish are presented in Table 6.8.

Table 6.7
link to table

Table 6.9
link to table

6.3.6 Estrogenic effects

Jobling et al. (1985) /100/ studied the estrogenic effects of a range of chemicals, including DBP, commonly found in sewage effluents. Using cytosolic extract from liver of rainbow trout, Oncorhynchus mykiss, in which estradiol receptor-binding sites are present in both female and male fish, Jobling et al. (1985) /100/ documented that DBP binds to the receptor, inhibiting the binding of natural estradiol. It has also been shown that DBP has mitogenic effect on the in vitro growth of human breast cancer cell (ZR-75) at test concentrations of 2.78 mg/l. In transiently transfected MCF 7 breast cancer cells, DBP was reported to affect the transcriptional activity of the estrogen receptor /100/. DBP concentrations in the range from 2.8 to 27.8 mg/l stimulated the activity.

In a study by Harris et al. (1997) /62/, DBP was found to have estrogenic activity using a recombinant yeast screen. The relative potency of DBP was approx. 1·10⁶ times less than 17b-estradiol. In addition, it was found that DBP at a concentration of 10^-11 M increases the transcriptional activity in the presence of natural 17b-estradiol.

Levels of 52-794 mg/kg DBP were daily dosed to male and female rats /101/. In tests for determination of the affected sex, the number of offspring was unchanged but the weights of pups from treated females were significantly decreased and offspring from treated males were unchanged. F₁ necropsy results revealed that epididymal sperm counts and testicular spermatid head counts were significantly decreased in the highest dose group. In conclusion, this study showed that DBP is a reproductive/developmental toxicant in Sprague-Dawley rats exposed both as adults and during development. It also indicates that the adverse reproductive/developmental effects of DBP on the second generation were greater than on the first generation.

Sharpe et al. (1995) /102/ assessed whether exposure of male rats to xenoestrogens during gestation and during the first three weeks after birth affects the size of their testes and sperm production in adult life. No effects of DBP were described. Likewise, Meek et al. (1997) /103/ measured changes in the reproductive organs of female rats. They showed that DBP produced no estrogenic effects. It has, however, to be noted that the conclusions made by Meek et al. /103/ are based on unpublished data.

6.3.7 Summary and conclusions

The effect concentrations found for different micro-organisms showed relatively high variability.

DBP has been shown acutely toxic (EC₅₀ or LC₅₀ values) to algae, crustaceans and fish in the range of 0.35-8.0 mg/l and is thus considered very toxic to aquatic organisms. In an estaurine microcosm, the abundance and diversity of crustaceans were affected at low concentrations and NOEC was determined to 0.04 mg/l.

NOEC levels in chronic toxicity tests with crustaceans and fish were both close to 0.1 mg/l. Compared to the NOEC levels derived in the acute toxicity tests, further toxicity was achieved in the long-term tests.

DBP has shown to be estrogenic in vitro, stimulating human breast cancer cell growth and transcriptional activity of the estrogen receptor. In some in vivo tests, DBP has shown to be testis toxic in adult rats, causing atrophy of the testes, prostate, seminal vesicles and epididymis while in other tests, no estrogenic effects could be detected.

DBP concentrations occasionally reach levels of up to 623 µg/l in river water. This worst case concentration has not the potential to exert estrogenic effects if tested on human breast cancer cells. Lack of in vivo estrogenic effect studies in the aquatic environment makes an assessment of the potential estrogenic effects of DBP in wildlife difficult.

6.4 Environmental hazard classification

DBP has been shown acutely toxic (EC₅₀ or LC₅₀ values) to algae, crustaceans and fish in the range of 0.35-8.0 mg/l.

NOEC levels in chronic toxicity tests with crustaceans and fish were both close to 0.1 mg/l. Compared to the NOEC levels derived in the acute toxicity tests further toxicity was achieved in the long-term tests.

DBP is readily biodegradable in standard laboratory tests and, hence, the substance is expected to be mineralized rapidly in the aerobic part of a sewage treatment plant. The studies of anaerobic biodegradability indicate a potential for mineralization of DBP during anaerobic treatment of sludge.

DBP is bioaccumulative in aquatic biota, which is demonstrated by experimentally derived BCF values of up to 2125 for fish.

The water solubility of DBP is = 10 mg/l which is well above the cut-off value of 1 mg/l.

Considering the criteria for environmental hazard classification (EEC 1993) and the above evaluation of the environmental fate and effect of Di-n-butyl phthalate it is proposed that DBP is classified "N; R50/53: Very toxic to aquatic organisms, may cause long-term adverse effects in the aquatic environment".

6.5 PNEC for the aquatic compartment

Toxicity data are available on short-term tests with bacteria, protozoa, algae, crustaceans and fish. All toxicity data are more or less in the same range with EC/LC₅₀ values from 0.35 mg/l to a few mg/l. Long-term toxicity data are available on algae, crustaceans and fish with NOEC values for crustaceans and fish at 0.1 mg/l as the lowest. A NOEC for effects on abundance and diversity in an estuarine microcosm was determined at 0.04 mg/l. An assessment factor of 10 for the lowest long-term NOEC should be used for deriving a PNEC_aquatic. However, considering the fact that the substance is bioaccumulative and has a potential estrogenic activity, an extra safety factor of 10 is used resulting in a proposed PNEC_aquatic » 0.001 mg/l.