Review of Environmental Fate and Effects of Selected Phthalate Esters

Contents

1 Introduction

2 Summary and conclusions
2.1 Concentrations in the environment
2.2 Degradation
2.3 Bioaccumulation
2.4 Toxicity
2.5 Estrogenic effects
2.6 Environmental Hazard Classification
2.7 Predicted No-Effect-Concentrations for the aquatic environment
2.8 Summary of the environmental fate and effect of 6 phthalate esters

3 Resumé og konklusioner
3.1 Koncentrationer i miljøet
3.2 Nedbrydning
3.3 Bioakkumulering
3.4 Toksicitet
3.5 Østrogenlignende effekter
3.6 Miljøfareklassifikation
3.7 Nul-effekt-koncentrationer (PNEC) for vandmiljøet
3.8 Resumé af 6 phthalatesteres miljømæssige skæbne og effekter

4 Dimethyl Phthalate (DMP)
4.1 Physico-chemical properties
4.1.1 Water solubility
4.1.2 Octanol-water partition coefficient
4.1.3 Summary
4.2 Environmental concentrations and fate
4.2.1 Concentrations in the environment
4.2.2 Abiotic degradation
4.2.3 Biodegradation
4.2.4 Bioaccumulation
4.2.5 Summary and conclusion
4.3 Effects
4.3.1 Toxicity to micro-organisms
4.3.2 Toxicity to algae
4.3.3 Toxicity to invertebrates
4.3.4 Toxicity to fish
4.3.5 Estrogenic effects
4.3.6 Summary and conclusions
4.4 Environmental hazard classification
4.5 PNEC for the aquatic compartment

5 Diethyl Phthalate (DEP)
5.1 Physico-chemical properties
5.1.1 Water solubility
5.1.2 Octanol-water partition coefficient
5.1.3 Summary
5.2 Environmental concentrations and fate
5.2.1 Concentrations in the environment
5.2.2 Abiotic degradation
5.2.3 Biodegradation
5.2.4 Bioaccumulation
5.2.5 Summary and conclusion
5.3 Effects
5.3.1 Toxicity to micro-organisms
5.3.2 Toxicity to algae
5.3.3 Toxicity to invertebrates
5.3.4 Toxicity to fish
5.3.5 Estrogenic effects
5.3.6 Summary and conclusions
5.4 Environmental hazard classification
5.5 PNEC for the aquatic environment

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

7 Butylbenzyl Phthalate (BBP)
7.1 Physico-chemical properties
7.1.1 Water solubility
7.1.2 Octanol-water partition coefficient
7.1.3 Summary
7.2 Environmental concentrations and fate
7.2.1 Concentrations in the environment
7.2.2 Abiotic degradation
7.2.3 Biodegradation
7.2.4 Bioaccumulation
7.2.5 Summary and conclusion
7.3 Effects
7.3.1 Toxicity to micro-organisms
7.3.2 Toxicity to algae
7.3.3 Toxicity to invertebrates
7.3.4 Toxicity to fish
7.3.5 Estrogenic effects
7.3.6 Summary and conclusions
7.4 Environmental hazard classification
7.5 PNEC for the aquatic compartment

8 Diisononyl Phthalate (DINP)
8.1 Physico-chemical properties
8.1.1 Water solubility
8.1.2 Octanol-water partition coefficient
8.1.3 Summary
8.2 Environmental concentrations and fate
8.2.1 Concentrations in the environment
8.2.2 Abiotic degradation
8.2.3 Biodegradation
8.2.4 Bioaccumulation
8.2.5 Summary and conclusion
8.3 Effects
8.3.1 Toxicity to micro-organisms
8.3.2 Toxicity to algae
8.3.3 Toxicity to invertebrates
8.3.4 Estrogenic effects
8.3.6 Summary and conclusions
8.4 Environmental hazard classification
8.5 PNEC for the aquatic compartment

9 Diisodecyl Phthalate (DIDP)
9.1 Physico-chemical properties
9.1.1 Water solubility
9.1.2 Octanol-water partition coefficient
9.1.3 Summary
9.2 Environmental concentrations and fate
9.2.1 Concentrations in the environment
9.2.2 Abiotic degradation
9.2.3 Biodegradation
9.2.4 Bioaccumulation
9.2.5 Summary and conclusion
9.3 Effects
9.3.1 Toxicity to micro-organisms
9.3.2 Toxicity to alga
9.3.3 Toxicity to invertebrates
9.3.4 Toxicity to fish
9.3.5 Estrogenic effects
9.3.6 Summary and conclusions
9.4 Environmental hazard classification
9.5 PNEC for the aquatic compartment

10 References

1 Introduction

The present review of the environmental fate and effects of Dimethyl Phthalate (DMP); Diethyl Phthalate (DEP); Di-n-butyl Phthalate (DBP); Butylbenzyl Phthalate (BBP); Diisononyl Phthalate (DINP) and Diisodecyl Phthalate (DIDP) has been prepared by VKI for the Danish Environmental Protection Agency.

The main objective of the present review was to collect existing knowledge about the above mentioned phthalate esters, mainly based on data from existing reviews and easily available handbooks and databases. Thus a thorough literature search has not been performed.

The following physico-chemical and ecotoxicological properties have been covered:

• water solubility
• octanol-water partition coefficient
• degradation
• bioaccumulation
• toxicity to aquatic organisms
• estrogenic effects

Based on the available data, an Environmental Hazard Classification has been proposed for each of the phthalate esters. Furthermore, Predicted No-Effect Concentrations for the aquatic environment (PNEC_aquatic) have been proposed according to the methods used in the EU and DK, i.e. the methods described in the EU Technical Guidance Document for risk assessment of chemical substances.
 

2 Summary and conclusions

2.1 Concentrations in the environment

The quality of chemical analysis of phthalates in environmental samples has been under debate during recent years. One of the main problems is the common use of plastic equipment in laboratories often containing platicizers. Consequently, the samples to be analysed may be contaminated during sampling, storage, processing as well as during analysis if a proper methodology has not been implemented. Thus, many of the results reported - especially in older references - may overestimate the concentrations in the samples due to contamination. This can be illustrated by measurements of e.g. DBP in rivers, in which the measured concentrations vary from 0.001 to 622.9 µg/l.

2.2 Degradation

Phthalate esters can undergo hydrolysis in two steps under production of mono-ester and a free alcohol in the first step and phthalic acid and a free alcohol in the second step. Hydrolysis, however, seems to play only an insignificant role for the degradation under natural environmental conditions with increasing hydrolysis rates at increasing pH /1, 2, 3/.

Photodegradation may be an important degradation pathway in atmosphere with predicted half-lives in the range of a few days /4/. However, in the soil and aquatic environments, the light intensity is so low that no significant photodegradation can be expected /1/.

In general, phthalate esters with short alkyl chain length are readily biodegradable but the mineralization rate decreases with increasing ester chain length.

Because of the ubiquitous use of phthalates, many sewage treatment plants contain adapted micro-organisms capable of degrading these substances. Also in anaerobic sewage sludge digesters, a potential for mineralization of some phthalate esters may be expected. Exceptions to this degradation behaviour may be the long-chain length phthalates, and only few data are available on the degradation of these substances under anaerobic conditions. However, relatively high amounts of long-chain length phthalates are found in sewage sludge demonstrating a low biodegradability under normal conditions in sewage treatment plants.

In tests performed at environmentally relevant conditions, mineralization of the short-chain phthalate esters has been found. The degradation of longer chain phthalate esters is lower and often only a primary biodegradation is found. Moreover, especially at low temperatures, the degradation is considerably slower than determined at the standardised laboratory conditions.

2.3 Bioaccumulation

In general, phthalate esters should be expected to be bioaccumulative due to their log K_ow values ranging from 1.61 (DMP) to >8 (DINP, DIDP). However, the accumulation is influenced by the capability of an organism to metabolize the substances. Numerous experiments have shown that bioaccumulation of phthalate esters in the aquatic and terrestrial food chains is limited due to biotransformation. The metabolic capability increases with increasing trophic levels.

Phthalate ester metabolism appears to depend upon both species and exposure route. Results indicate that mean BCFs are highest for algae and lowest for fish with invertebrates exhibiting intermediate values /1/. These findings are consistent with previous studies by Wofford et al. (1981) /5/ who found that the extent of phthalate ester biotransformation increased as follows: molluscs < crustaceans < fish.

2.4 Toxicity

Detailed studies of the mode of toxic action of phthalate esters in aquatic organisms are lacking, however, polar narcosis is generally accepted as being the primary mode of action.

The low water solubility of some phthalate esters causes problems when exposing aquatic organisms in toxicity tests. The formation of micro droplets, surface films and adsorption to surfaces lead to difficulties in maintaining steady exposure concentrations and/or cause direct physical interference. The low water solubility has led to the widespread use of carrier solvents in toxicity testing. Reported aqueous effect concentrations often greatly exceeds true water solubilities in tests performed with higher molecular weight phthalate esters. Water solubility, biodegradation, and sorption may thus significantly influence the results of these aquatic toxicity tests.

When test solutions are prepared in concentrations higher than 'true' water solubility of the phthalate esters, an emulsion of microdroplets consisting of pure chemical may be formed. The formation of microdroplets or surface films may contribute to possible effects by direct physical interference.

Small colloids may increase the apparent water solubility by sorbing lipophilic substances. They may, however, either decrease or increase the bioavailability and thus the toxicity. For most substances, the presence of particulates and colloids probably decreases the bioavailability, but for certain types of organisms (especially suspension feeders and detritovores) the reverse effect might be the case.

When interpreting toxicity data on sparingly soluble substances such as especially the higher molecular weight phthalate esters (DINP and DIDP), it is very important that the above parameters are taken into account.

2.5 Estrogenic effects

The potential for estrogenic effects in wildlife has been evaluated by means of extrapolation from in vitro and in vivo studies with rats. Lack of in vivo studies for phthalate esters in aquatic environments makes assessment of their potential estrogenic effects in aquatic wildlife difficult.

2.6 Environmental Hazard Classification

The phthalate esters have not been considered for environmental hazard classification by the EU "Labelling Group". However, based on the present review of the environmental fate and effects of the phthalate esters, classification proposals have been derived.

The main parameters to be considered for the environmental hazard classification (EEC 1993) are:

• acute toxicity to algae, daphnia and fish
• chronic toxicity to daphnia and fish
• ready biodegradability
• bioaccumulation potential (log K_ow or experimentally derived BCF)
• water solubility

2.7 Predicted No-Effect-Concentrations for the aquatic environment

Predicted No Effect Concentrations (PNEC_aquatic) have been derived for the phthalate esters evaluated. When possible, the PNEC values were derived from long-term studies on algae, crustaceans and fish. As, however, only short-term studies exist for one of the phthalate esters evaluated, PNEC have been derived from these. For DINP and DIDP, no PNEC_aquatic could be derived.

2.8 Summary of the environmental fate and effect of 6 phthalate esters

The fate and effect of Dimethyl Phthalate (DMP); Diethyl Phthalate (DEP); Di-n-butyl Phthalate (DBP); Butylbenzyl Phthalate (BBP); Diisononyl Phthalate (DINP) and Diisodecyl Phthalate (DIDP) have been summarized in Table 2.1.

Table 2.1
Summary of the environmental fate and effect of 6 phthalate esters

N.d.: Not derived.
 

3 Resumé og konklusioner

3.1 Koncentrationer i miljøet

Kvaliteten af de kemiske analyser af phthalater i miljøprøver, f.eks. jord-, vand-, slam- og spildevandsprøver, har været diskuteret i de senere år. Et af hovedproblemerne er laboratoriernes anvendelse plastikudstyr, som ofte indeholder blødgørere. De prøver, som skal analyseres, kan således risikere at blive forurenede både ved prøveudtagning, under opbevaring, ved behandling og ved selve analysen, hvis der ikke udvises tilstrækkelig forsigtighed. Mange af de forsøgsresultater, som er blevet afrapporteret, især i ældre referencer, kan således have overvurderet koncentrationerne i prøverne på grund af en sådan forurening. Dette kan f.eks. illustreres med målinger af DBP i floder, hvor de målte koncentrationer varierer fra 0,001 til 622,9 µg/l.

3.2 Nedbrydning

Phthalatestre kan hydrolysere i to trin med dannelse af en monoester og en fri alkohol i første trin og phthalsyre og fri alkohol i andet trin. Hydrolyse synes imidlertid ikke at spille nogen særlig betydningsfuld rolle for nedbrydningen under naturlige miljømæssige forhold, hvor hydrolysehastigheden stiger i takt med stigende pH /1, 2, 3/.

Fotolytisk nedbrydning kan være et vigtigt nedbrydningsforløb i atmosfæren med beregnede halveringstider på bare nogle få dage /4/. I jord- og vandmiljøet er lysintensiteten imidlertid så lav, at der ikke kan forventes nogen betydende fotolytisk nedbrydning /1/.

Phthalatestere med kort alkyl-kædelængde er generelt let bionedbrydelige, men nedbrydningshastigheden er lavere for stoffer, hvor esterkæderne er længere.

På grund af den meget udbredte brug af phthalater indeholder mange renseanlæg i dag adapterede mikroorganismer, som kan nedbryde disse stoffer. Desuden kan der forventes en potentiel nedbrydning af nogle phthalater i anaerobe rådnetanke. Phthalater med lang kædelængde kan dog være undtagelser fra dette nedbrydningsforløb, og der findes kun få data om nedbrydningen af disse stoffer under anaerobe forhold. Der er imidlertid fundet relativt store mængder af lang-kædede phthalater i spildevandsslam, hvilket viser, at bionedbrydeligheden er lav under normale forhold i et renseanlæg.

Der er observeret mineralisering af kort-kædede phthalatestere i forsøg udført under miljømæssigt relevante forhold. Nedbrydningsgraden er lavere for phthalater, hvor esterkæderne er længere, og ofte er der kun observeret primær bionedbrydning. Nedbrydningshastigheden er desuden betydeligt langsommere, især ved lave temperaturer, end den som er bestemt under standardiserede laboratorieforhold.

3.3. Bioakkumulering

Phthalatestere forventes generelt at være bioakkumulerbare på grund af deres log K_ow værdier, som ligger mellem 1,61 (DMP) og >8 (DINP, DIDP). Bioakkumuleringen påvirkes imidlertid af organismens evne til at metabolisere stofferne. Adskillige forsøg har vist, at bioakkumuleringen af phthalatestere i både de akvatiske og de terrestriske fødekæder er begrænset på grund af den biologiske omdannelse. Evnen til at metabolisere stofferne øges i takt med stigende trofisk niveau.

Det ser ud til, at phthalatestermetabolismen både afhænger af arten og eksponeringsvejen. Resultaterne indikerer, at de gennemsnitlige BCF-værdier er højest for alger og lavest for fisk, mens invertebrater udviser mellemliggende værdier /1/. Disse resultater er i overensstemmelse med tidligere undersøgelser udført af Wofford et al. (1981) /5/, som har observeret, at omfanget af metabolismen af phthalatestere stiger som følger: bløddyr < krebsdyr < fisk.

3.4 Toksicitet

Der findes ingen detaljerede undersøgelser af phthalatesteres toksiske virkemekanisme i akvatiske organismer. Det er dog almindeligt accepteret, at polær narkose er den primære virkemekanisme.

Nogle phthalatesteres lave vandopløslighed kan give problemer ved eksponeringen af akvatiske organismer i toksicitetstest. Dannelsen af mikrodråber, overfladefilm og adsorption til overflader gør det besværligt at opretholde stabile testkoncentrationer og/eller medfører direkte fysisk indvirkning. Den lave vandopløselighed har ført til, at brugen af opløsningsmidler er almindeligt udbredt inden for toksicitetstest. Refererede vandige effektkoncentrationer ligger ofte langt over den virkelige vandopløselighed i test, som er udført med phthalatestere med høj molekylvægt. Vandopløselighed, biologisk nedbrydning og sorption kan således påvirke resultaterne fra disse akvatiske toksicitetstest markant.

Der kan dannes en emulsion af mikrodråber, som består af uopløst kemikalie, når testopløsningerne er lavet i højere koncentrationer end phthalatesternes "ægte" vandopløselighed. Dannelsen af mikrodråber eller overfladefilm kan bidrage til mulige effekter ved direkte fysisk indvirkning.

Små kolloider kan forhøje den tilsyneladende vandopløselighed ved at sorbere til lipofile stoffer. De kan imidlertid enten nedsætte eller forøge biotilgængeligheden og dermed toksiciteten. For de fleste stoffers vedkommende nedsætter tilstedeværelsen af partikler og kolloider sandsynligvis biotilgængeligheden, men for visse typer organismer (især filtratorer og sedimentædere) kan det modsatte være tilfældet.

Det er meget vigtigt at tage ovennævnte parametre i betragtning, når der benyttes toksicitetsdata for svært opløselige stoffer, som f.eks. phthalatestere med høj molekylvægt (DINP og DIDP).

3.5 Østrogenlignende effekter

De potentielle østrogenlignende effekter på dyrelivet er blevet vurderet ved hjælp af ekstrapolation fra in vitro og in vivo forsøg med rotter. Det er svært at vurdere phthalatesteres potentielle østrogenlignede effekter på vildtlevende akvatiske organismer, da der ikke findes in vivo forsøg med phthalatestere i vandmiljøet.

3.6 Miljøfareklassifikation

Phthalatesteres miljøfareklassifikation er ikke blevet vurderet af EUs "Mærkningsgruppe". Der er imidlertid blevet udarbejdet nogle klassificeringsforslag ud fra nærværende gennemgang af phthalatesternes miljømæssige skæbne og effekter.

De vigtigste parametre i forbindelse med miljøfareklassifikationen (EEC 1993) er følgende:

• akut toksicitet over for alger, dafnier og fisk
• kronisk toksicitet over for dafnier og fisk
• let bionedbrydelighed
• bioakkumuleringspotentiale (log K_ow eller BCF-værdier opnået ved forsøg)
• vandopløselighed

3.7 Nul-effekt-koncentrationer (PNEC) for vandmiljøet

Der er udarbejdet nul-effekt-koncentrationer (PNEC_aquatic) for de vurderede phthalatestere. PNEC-værdierne er så vidt muligt udledt ud fra langtidsforsøg med alger, krebsdyr og fisk. For en af phthalatestrene findes der dog kun korttidsforsøg, og PNEC-værdierne er derfor udarbejdet fra disse. Det har ikke været muligt at udarbejde PNEC-værdier for DINP og DIDP.

3.8 Resumé af 6 phthalatesteres miljømæssige skæbne og effekter

Tabel 3.1 giver en oversigt over den miljømæssige skæbne og effekt af dimethylphthalat (DMP), diethylphthalat (DEP), di-n-butylphthalat (DBP), butylbenzylphthalat (BBP), diisononylphthalat (DINP) og diisodecylphthalat (DIDP).

Tabel 3.1
Oversigt over 6 phthalatesteres miljømæssige skæbne og effekter

i.u.: Ikke udledt
 

4 Dimethyl Phthalate (DMP)

DMP is used as a plasticizer in latex, cellulose acetate film and plastics. DMP is as a constituent of rubber, coating agents, safety glass, moulding powders, insect repellents and perfumes. DMP leaches to the environment from tubings, dishes, paper, containers by general use of plastics and the above listed products /6/.

4.1 Physico-chemical properties

DMP (C₁₀H₁₀O₄), CAS No.: 131-11-3, with an alkyl chain length of 1,1 /1/ is a colourless liquid. The molecular weight is 194.2 g/mol. DMP has a melting point of 0°C and a boiling point at 282°C /6/. The density is 1.192 g/ml and the vapour pressure is 2·10^-3 mmHg at 25°C /1/.

4.1.1 Water solubility

DMP is a low molecular weight phthalate. Several aqueous solubility data on DMP are referred to in the literature. Independent experimental measurements are generally in good agreement and believed to be reliable for lower molecular weight molecules. Literature values range from 2810 to 4320 mg/l. The solubility has been calculated to 2179 mg/l /7/. In a literature review by Staples et al. /1/, it was concluded that a water solubility of about 4200 mg/l was the most likely value based on available evidence.

4.1.2 Octanol-water partition coefficient

Fairly consistent log K_ow values are seen in the literature for lower molecular weight phthalates as DMP. Reported log K_ow values are only ranging from 1.46 to 1.90. A log K_ow value of 1.56 has been calculated /7/. According to Staples et al. /1/, the most likely log K_ow value based on available evidence was concluded to be 1.61.

4.1.3 Summary

The physico-chemical properties on DMP are summarized in Table 4.1.

Table 4.1
Physico-chemical properties of Dimethyl Phthalate (DMP)

4.2 Environmental concentrations and fate

4.2.1 Concentrations in the environment

The content of DMP in wastewater and sewage sludge from Danish treatment plants has been measured at one occasion during recent years. An overview of the results is given in Table 4.2.

Table 4.2
Dimetyl Phthalate (DMP) in treatment plants

1) Cited from /10/.

As no data were available on measurements in inlet and outlet waste water, it was not possible to derive any mass balances.

No data available.

4.2.2 Abiotic degradation

Wolfe et al. (1980) /2/ measured the hydrolysis rate constant of DMP and estimated a half-life of 3.2 years at alkaline conditions. The hydrolysis half-life at neutral pH and 25°C is estimated to 2.7 years /7/.

No experimental data on photodegradation of DMP are available. Estimated photodegradation half-lives in the atmosphere are in the range from 9.3 to 93 days /4, 7/. DMP in pure water is photodegraded by irradiation with UV light with a half-life of 13 hours /11/. However, in the aquatic environment only insignificant photodegradation is expected /1/.

4.2.3 Biodegradation

The ready biodegradability of DMP was determined in the OECD 301C test resulting in a degradation of 90-98% /12/.

Staples et al. (1996) /1/ refer to an investigation by Changming & Kang (1990) /13/ showing a degradation of 99.6% after 4 days incubation.

Sugatt et al. (1984) /14/ using an acclimated inoculum demonstrated a biodegradability of DMP of 86% after 28 days. Staples et al. (1996) /1/ refer to a study of Aichinger et al. (1992) /15/ using an acclimated inoculum demonstrating a degradability of 96%.

Staples et al. (1996) /1/ have reviewed the biodegradability of DMP and referred numerous studies showing a high degree of primary biodegradability - in general between 90% and 100%.

Howard (1989) /11/ refers to studies on the biodegradation of DMP in sewage treatment plants demonstrating that the total removal frequently approached 100% while mineralization ranged between 58 and 88%.

Staples et al. (1996) /1/ refer to studies showing anaerobic primary biodegradation at 30-37°C in the range from 18% to 100% and anaerobic ultimate biodegradation in the range from 41% to 100%.

Hattori et al. (1975) /16/ demonstrated 100% primary degradation after 8 days in freshwater in a river die-away test, but only 0-32% after 7-14 days in marine waters.

4.2.4 Bioaccumulation

For DMP, only one bioaccumulation study performed with fish was found. A total BCF of 57 for Bluegill Sunfish (Lepomis macrochirus) was reported by Barrows et al. (1980) /17/ (exposure concentration: 8.7 µg/l; test procedure: flow through; exposure period: not known). The low bioaccumulation potential is in conformity with the log K_ow » 1.6.

4.2.5 Summary and conclusion

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

DMP 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 DMP during anaerobic treatment of sludge.

DMP has a low bioaccumulation potential demonstrated by both log K_ow » 1.6 and the experimentally derived BCF value of 57 for fish.

4.3 Effects

4.3.1 Toxicity to micro-organisms

The toxicity studies with micro-organisms are summarized in Table 4.3. The table contains data on both bacteria and protozoa.

Table 4.3
Toxicity of Dimethyl phthalate to micro-organisms

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

From the above results, DMP seems to have relatively low toxicity to micro-organisms.

4.3.2 Toxicity to algae

The short-term toxicity studies with DMP for freshwater and marine algae are summarized in Table 4.4.

Table 4.4
Toxicity of Dimethyl phthalate to algae

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

The toxicity data obtained on the different algae species seem to be in close agreement, except for the tests with Chlorella pyrenoidosa and Gymnodinium breve, which show higher effect concentrations. The relatively high EC₅₀ value (142 mg/l) in the 6-day test Selenastrum capricornutum can be attributed to experimental problems during the relatively long exposure period or to the biodegradability of the substance.

4.3.3 Toxicity to invertebrates

The short-term toxicity data on DMP to freshwater and marine invertebrates are presented in Table 4.5 and the long-term toxicity data on DMP to freshwater and marine invertebrates are presented in Table 4.6.

Table 4.5
Short-term toxicity of Dimethyl phthalate to aquatic invertebrates

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

Table 4.6
Long-term toxicity of Dimethyl phthalate to aquatic invertebrates

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

The toxicity data obtained with the different crustacean species are in close agreement except for the test with Paratanytarsus parthenogenica.

4.3.3.Toxicity to fish

The short-term toxicity data on DMP to freshwater and marine fish are presented in Table 4.7 and the long-term toxicity data on DMP to freshwater and marine fish species are presented in Table 4.8.

Table 4.7
Short-term toxicity data on Dimethyl phthalate to fish

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

Table 4.8
Long-term toxicity data on Dimethyl phthalate to fish

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

From the results obtained by Springborn Bionomics (1987) /43/ with Pimephales promelas, it can be seen that no further toxicity is obtained when the exposure period is increased from 96 h to 144 h. This indicates that steady state conditions and thus the maximum toxicity of DMP are reached during a 96 h test period.

4.3.5 Estrogenic effects

No data are available.

4.3.6 Summary and conclusions

DMP seems to have a relatively low toxicity to micro-organisms.

DMP has been shown acutely toxic (EC₅₀ or LC₅₀ values) to algae, crustaceans and fish in the range 26-377 mg/l, however, with most values in the range 25-50 mg/l irrespective of the trophic level investigated. DMP is thus, solely based on aquatic toxicity data, considered harmful to aquatic organisms.

NOEC levels in chronic toxicity tests with crustaceans and fish were both close to 10 mg/l. Compared to the NOEC levels derived in the acute toxicity tests no further toxicity was achieved in the long-term tests.

No data are available in which the estrogenic effect of DMP has been evaluated.

4.4 Environmental hazard classification

DMP has been shown acutely toxic (EC₅₀ or LC₅₀ values) to algae, crustaceans and fish in the range 26-377 mg/l, however, with most values in the range 25-50 mg/l irrespective of the trophic level investigated.

NOEC levels in chronic toxicity tests with algae, crustaceans and fish were all close to 10 mg/l. Compared to the NOEC levels derived in the acute toxicity tests no further toxicity was achieved in the long-term tests.

DMP 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 DMP during anaerobic treatment of sludge.

DMP is bioaccumulative in aquatic biota, which is demonstrated by the experimentally derived BCF value of 57 for fish.

The water solubility of DMP is » 4200 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 effects of Dimethyl phthalate, it is proposed that DMP should not be classified as dangerous to the aquatic environment.

4.5 PNEC for the aquatic compartment

Long-term NOECs for three trophic levels are available, all of them in the same concentration range. The lowest NOEC available is the 21 d NOEC for Daphnia magna at 9.6 mg/l. Considering the fact that the substance is readily biodegradable and with a low bioaccumulation potential, an assessment factor of 10 is proposed for the derivation of a PNEC_aquatic resulting in a proposed PNEC_aquatic = 1 mg/l.
 

5 Diethyl Phthalate (DEP)

DEP is used as a plasicizer in plastics, food packaging application. DEP is a dye application agent, and a diluent in polysulfide dental impression materials solvent; wetting agent; camphor substitute; used in perfumery; alcohol denaturant and as a component in insecticidal sprays /6/.

5.1 Physico-chemical properties

DEP (C₁₂H₁₄O₄), CAS No.: 84-66-2, with an alkyl chain length of 2,2 /1/ is a clear, stable, odourless liquid. The molecular weight is 222.2 g/mol. DEP has a melting point of about -40.5°C and a boiling point at 298°C. The density is 1.118 g/ml and the vapour pressure is 1·10^-3 mmHg at 25°C /1/.

5.1.1 Water solubility

DEP is a low molecular weight phthalate. Several aqueous solubility data on DEP are referred to in the ranging from 400 to 7028 mg/l. The solubility has been calculated to 260.3 mg/l /7/. In a literature review by Staples et al. /1/, it was concluded that a water solubility of about 1100 mg/l was the most likely value based on available evidence.

5.1.2 Octanol-water partition coefficient

Reported log K_ow values are ranging from 2.21 to 3.00. A log K_ow of 2.47 has been calculated /7/. According to Staples et al. /1/, the most likely log K_ow value based on available evidence was concluded to be 2.38.

5.1.3 Summary

The physico-chemical properties on DEP are summarized in Table 5.1.

Table 5.1
Physico-chemical properties of Diethyl Phthalate (DEP)

5.2 Environmental concentrations and fate

5.2.1 Concentrations in the environment

The content of DEP in wastewater and sewage sludge from Danish treatment plants has been measured at one occasion during recent years. An overview of the results is given in Table 5.2.

Table 5.2
Diethyl Phthalate (DEP) in treatment plants

1) Cited from /10/.

As no data were available on measurements in inlet and outlet waste water, it is not possible to derive any mass balances.

No data are available.

5.2.2 Abiotic degradation

Wolfe et al. (1980) /2/ measured the hydrolysis rate constant of DEP and estimated a half-life of 8.8 years at pH 7. The hydrolysis half-life at neutral pH and 25°C range is estimated to 2.9 years /7/.

No experimental data on photodegradation of DEP are available. Estimated photodegradation half-lives in the atmosphere are in the range from 1.8 to 18 days /4, 7/. In the aquatic environment only insignificant photodegradation is expected /1/.

5.2.3 Biodegradation

Staples et al. (1996) /1/ refer to an investigation by Changming & Kang (1990) /13/ showing a degradation of 99.2% after 6 days of incubation.

By use of an acclimated inoculum, Sugatt et al. (1984) /14/ found a biodegradability of 95% of DEP after 28 days of incubation. Staples et al. (1996) /1/ refer to a study of Aichinger et al. (1992) /15/ using acclimated inocula demonstrating a degradability of 93%.

Staples et al. (1996) /1/ have reviewed the biodegradability of DEP and referred to numerous studies showing a high degree of primary biodegradability - in general between 90% and 100%.

In a simulation of a biological sewage treatment plant, Patterson & Kodukala (1981) /47/ determined 79-98% removal of DEP.

Staples et al. (1996) /1/ refer to a few tests on anaerobic biodegradability of DEP showing a primary biodegradability of 64% to 100% and an ultimate biodegradability of 0% to 76% at 35-37°C and incubation for up to 70 days.

Hattori et al. (1975) /16/ demonstrated a 100% primary degradability of DEP after 6 days in freshwater in a river die-away test, and 14-68% after 14 days in marine waters.

5.2.4 Bioaccumulation

For DEP, only one bioaccumulation study performed with fish was found. A total BCF of 117 for Bluegill Sunfish (Lepomis macrochirus) was reported by Barrows et al. (1980) /17/ in a flow-through experiment and feeding ad libitum. Exposure concentration: 8.7 µg/l. Exposure period is not known.

5.2.5 Summary and conclusion

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

DEP 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 DEP during anaerobic treatment of sludge.

DEP has a moderate bioaccumulation potential demonstrated by both log K_ow » 2.4 and the experimentally derived BCF value of 117 for fish.

5.3 Effects

5.3.1 Toxicity to micro-organisms

The toxicity studies with micro-organisms are summarized in Table 5.3. The table contains data on both bacteria and protozoa.

Table 5.3
Toxicity of Diethyl phthalate to micro-organisms

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration. TT: Toxicity Threshold = geometric mean of NOEC and LOEC.

From the above results, DEP seems to have relatively low toxicity to micro-organisms.

5.3.2 Toxicity to algae

The toxicity studies with DEP for freshwater and marine algae are summarized in Table 5.4.

Table 5.4
Toxicity of Diethyl phthalate to algae

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

The toxicity data obtained on the different algae species seem to be in close agreement except the 96 h test on effects on the growth rate of Gymnodinium breve where a relatively low EC₅₀ value was found.

5.3.3 Toxicity to invertebrates

The short-term toxicity data on DEP to freshwater and marine invertebrates are presented in Table 5.5 and the long-term toxicity data on DEP to freshwater and marine invertebrates are presented in Table 5.6.

Table 5.5
Short-term toxicity of Diethyl phthalate to aquatic invertebrates

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

Table 5.6
Long-term toxicity of Diethyl phthalate to aquatic invertebrates

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

As seen from the tables, NOEC values of 13 mg/l and 25 mg/l, respectively, were obtained in 21 d reproduction tests with Daphnia magna. However, a lower NOEC value was obtained in a 96 h mortality test with the marine crustacean Mysidopsis bahia (2.7 mg/l, measured concentration).

5.3.4 Toxicity to fish

The short-term toxicity data on DEP to freshwater and marine fish are presented in Table 5.7. No long-term toxicity studies with fish were found for DEP.

Table 5.7
Short-term toxicity data on Diethyl phthalate to fish

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

5.3.5 Estrogenic effects

In an investigation by Harris et al. (1997) /62/, DEP was shown to have weak estrogenic activity in an in vitro recombinant yeast screen test, with a relative potency of approx. 5·10⁷ times less than 17b-estradiol.

5.3.6 Summary and conclusions

DEP has been shown acutely toxic (EC₅₀ or LC₅₀ values) to algae, crustaceans and fish in the range 7.6-131 mg/l, however, with most values in the range 20-40 mg/l irrespective of the trophic level investigated. DEP is thus, solely based on aquatic toxicity data, considered harmful to aquatic organisms. The NOEC values obtained in crustaceans and fish acute toxicity studies range from 1.7-46 mg/l. In long-term studies with algae and Daphnia magna NOEC-values of 3.7-25 mg/l were determined.

In an investigation by Harris et al. (1997) /62/, DEP was shown to have weak estrogenic activity in an in vitro recombinant yeast screen test, with a relative potency of approx. 5·10⁷ times less than 17b-estradiol.

5.4 Environmental hazard classification

DEP has been shown acutely toxic (EC₅₀ or LC₅₀ values) to algae, crustaceans and fish in the range 7.6-131 mg/l. The marine crustacean Mysidopsis bahia seems considerably more sensitive than Daphnia magna.

NOEC levels in chronic toxicity tests with algae were in the range 3.7-25 mg/l. Compared to the NOEC levels derived in the acute toxicity tests with crustaceans and fish, no further toxicity was achieved in these long-term tests.

DEP 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 DEP during anaerobic treatment of sludge.

DEP seems to have relatively low toxicity to micro-organisms.

DEP has a moderate bioaccumulation potential demonstrated by both log K_ow » 2.4 and the experimentally derived BCF value of 117 for fish.

The water solubility of DEP is » 1000 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 Diethyl phthalate, it is proposed that DEP should not be classified as dangerous for the aquatic environment.

5.5 PNEC for the aquatic environment

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 an EC₅₀ for algae at 3.0 mg/l and NOEC for crustaceans at 2.7 mg/l as the lowest short-term toxicity values. Long-term toxicity data are available on algae and crustaceans with a NOEC for algae at 3.7 mg/l as the lowest. An assessment factor of 50 for the two long-term NOEC should be used for deriving a PNEC_aquatic and considering the fact that the substance is readily biodegradable but has a moderate bioaccumulation potential and has a weak estrogenic activity, a PNEC_aquatic = 0.01 mg/l is proposed.
 

6 Di-n-butyl Phthalate (DBP)

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
Dibutyl Phthalate (DBP) in treatment plants

1) Cited from /3/.

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 µg/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
Toxicity of Dibutyl phthalate to micro-organisms

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

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
Toxicity of Dibuthyl phthalate to algae

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

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
Short-term toxicity of Dibutyl phthalate to aquatic invertebrates

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

Table 6.6
Long-term toxicity of Dibutyl phthalate to aquatic invertebrates

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

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
Short-term toxicity data on Dibuthyl phthalate to fish

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

Table 6.9
Long-term toxicity data on Dibuthyl phthalate to fish

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

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.
 

7 Butylbenzyl Phthalate (BBP)

BBP is used to plasticize or flexibilize synthetic resins, mainly polyvinylchloride /6/.

7.1 Physico-chemical properties

BBP (C19H20O4), CAS No.: 85-68-7, with a alkyl chain length of 4,7 (aryl) /1/ is a colourless oily liquid. The molecular weight is 312.4 g/mol. BBP has a melting point of about -35°C and a boiling point at 195-205°C /6/. The density is 1.111 g/ml at 25°C and the vapour pressure is 5.0·10^-6 mmHg at 25°C /1/.

7.1.1 Water solubility

For BBP, several aqueous solubility data are referred to in the literature ranging from 0.70 to 40.2 mg/l. The water solubility has been calculated to 0.67 mg/l /7/. In a literature review by Staples et al. /1/, it was concluded that a water solubility of about 2.7 mg/l was the most likely value based on available evidence.

7.1.2 Octanol-water partition coefficient

As for solubility, there are several different values in the literature for the octanol-water partition coefficient K_ow, differing with a factor of 22. Reported log K_ow values for BBP range from 3.57 to 4.91. A log K_ow of 4.91 has been calculated /7/. According to Staples et al. /1/, a log K_ow of 4.59 is the most likely value based on available evidence.

7.1.3 Summary

The approximate physico-chemical properties on BBP are summarized in Table 7.1.

Table 7.1
Physico-chemical properties of Butylbenzyl Phthalate (BBP)

7.2 Environmental concentrations and fate

7.2.1 Concentrations in the environment

The content of BBP in waste water 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 7.2.

Table 7.2
Butyl Benzyl Phthalate (BBP) in treatment plants

1) Cited from /3/.

It is a general picture that a high removal of BBP from the wastewater is found during the wastewater treatment. Mass balances have shown that only about 0.3% of the amount in the inlet water is found in sludge /9/, which indicates that a considerable amount of BBP may be reduced by anaerobic degradation.

A large number of data on concentrations of BBP in the environment is given in /3/ and only a brief overview will be given here.

Concentrations of BBP in 31 American rivers, lakes and estuaries are in the ranges from 0.2 to 2.4 µg/l, 0.35 to 0.45 µg/l and 0.3 µg/l, respectively. In surface water in German rivers and their main effluents, BBP was detected in concentrations of up to 3.4 µg/l and 49 µg/l, respectively.

Concentrations of BBP have been detected in river and lake sediments in the ranges from 60 to 14,000 µg/kg and 400 to 420 µg/kg, respectively. No information was found concerning BBP levels in marine sediments.

Soil sampled in the neighbourhood of phthalate-emitting plants contained concentrations of BBP of up to 100 µg/kg dry matter.

BBP has been detected in concentrations from 2.25 to 9.0 ng m^-3 in the city of Barcelona, Spain.

In biota, BBP has been detected at concentrations of up to 39 µg/kg wet weight (fish) and of up to 1.3 µg/kg dry matter (terrestrial plants).

7.2.2 Abiotic degradation

Wolfe et al. (1980) /2/ measured the hydrolysis rate constant of BBP and estimated a half-life of >0.3 years at alkaline conditions. The hydrolysis half-life at neutral pH and 25°C range is estimated to 1.4 years /7/.

No experimental data on photodegradation of BBP in the atmosphere are available. Estimated photodegradation half-lives are in the range from 0.5 to 5 days /4, 7/. Photodegradation of BBP in a 1 mg/l aqueous solution was studied by Gledhill et al. (1980) /104/ who reported less than 5% degradation in 28 days. Thus, in the aquatic environment only insignificant photodegradation is expected /1, 3/.

7.2.3 Biodegradation

The ready biodegradability of BBP was determined in the OECD 301C test resulting in a degradation of 81% after 14 days of incubation /12/. Staples et al. (1996) /1/ have reviewed the biodegradability of BBP and refer to three studies showing an ultimate biodegradability in the range from 10% to 65% after 28-30 days of incubation.

Saeger & Tucker (1976) /105/ tested the ultimate biodegradability during 27 days by incubating BBP with sludge, which had been acclimated for 14 days. 96% biodegradation was determined. Sugatt et al. (1984) /14/ using an acclimated inoculum demonstrated a biodegradability of BBP of 43% after 28 days of exposure. Györkös (1996) /3/ refers to an investigation by Gledhill et al. (1980) /104/ showing 96% biodegradation after 28 days by using an acclimated inoculum. Staples et al. (1996) /1/ report that the mineralization of BBP ranged from 66% to 96% after 27-28 days by use of acclimated inocula.

Saeger & Tucker (1976) /105/ determined the primary biodegradation of BBP in river water samples and found a rapid degradation of about 80% after 2 days of incubation.

In a semi-continuous activated sludge (SCAS) test, Saeger & Tucker (1976) /105/ determined a primary degradation of 93-99% of BBP after 24 hours of incubation.

Howard (1989) /11/ refers to an investigation demonstrating more than 90% degradation of BBP in about a week. Györkös (1996) /3/ refers to investigations showing 63% degradation after 1 week and more than 90% after 40 days under anaerobic conditions /67/, 50% after 29 days of incubation /106/, and 97% during anaerobic digestion /107/. Furthermore, Györkös (1996) /3/ refers to 78% and 88% degradation of BBP in anaerobic salt-marsh and fresh-water sediments after 22 and 35 days, respectively, of incubation in a study by Painter & Jones (1990) /106/. Staples et al. (1996) /1/ refer to studies showing anaerobic primary biodegradation at 30-37°C in the range from 50% to 100% for 7-100 days and anaerobic ultimate biodegradation in the range from 0% to 100% for 28-70 days.

Howard (1989) /11/ refers to an investigation demonstrating more than 95% primary degradation of BBP in a lake water microcosm after 7 days and 51-65% mineralization after 28 days. Györkös (1996) /3/ refers to investigations showing 80% primary degradation in unacclimated river water /105/ and 100% primary degradation in river water within 9 days /104/.

7.2.4 Bioaccumulation

In the following, only bioaccumulation studies on fish will be referred to as no studies on algae, crustaceans and insects were found.

Several bioaccumulation studies have been performed on Bluegill sunfish (Lepomis macrochirus) with total BCFs varying from 188 (17 days of exposure with a flow through test procedure and an exposure concentration of 2 µg/l) /126/ to 663 (flow through test procedure and an exposure concentration of 9.7 µg/l; exposure period is not known) /17/. The total BCF was by Carr et al. (1992) /64/ determined to 449 (3 days of exposure with a flow through test procedure and an exposure concentration of 34 µg/l), the corresponding BCF of the parent compound was by Staple et al. (1996) /1/ calculated to 12.

7.2.5 Summary and conclusion

BBP seems to be efficiently removed from waste waters based on the low output 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 BBP measured in some sludge samples.

BBP has been detected in soil and water (rivers, lakes and estuaries) at variable concentrations. In the aquatic environment, BBP concentrations of up to 49 µg/l have been found. In sediments BBP has been detected in concentrations of up to 14,000 µg/kg and in biota BBP has been detected at concentrations of up to 39 µg/kg wet weight (fish) and up to 1,256 µg/kg dry matter (terrestrial plants).

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

BBP 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 BBP during anaerobic treatment of sludge.

BBP is bioaccumulative in aquatic biota, which is demonstrated by the experimentally derived BCF value of up to 663 for fish.

7.3 Effects

7.3.1 Toxicity to micro-organisms

The toxicity studies with micro-organisms are summarized in Table 7.3. The table contains data on both bacteria and protozoa.

Table 7.3
Toxicity of Butylbenzyl phthalate to micro-organisms

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

7.3.2 Toxicity to algae

The toxicity studies with BBP for freshwater and marine algae are summarized in Table 7.4.

Table 7.4
Toxicity of Butylbenzyl phthalate to algae

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

The toxicity data obtained on the different algae species seem to be in close agreement.

7.3.3 Toxicity to invertebrates

The short-term toxicity data on BBP to freshwater and marine invertebrates are presented in Table 7.5. and the long-term toxicity data on BBP to freshwater and marine invertebrates are presented in Table 7.6.

Table 7.5
Short-term toxicity of Butylbenzyl phthalate to aquatic invertebrates

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

Table 7.6
Long-term toxicity of Butylbenzyl phthalate to aquatic invertebrates

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

In all tests performed, except one acute test with Mysidopsis bahia, the effect concentrations of BBP were at concentrations equal to or below the water solubility of the substance, furthermore, many of the effect concentrations determined are based on measured exposure concentrations.

7.3.4 Toxicity to fish

The short-term toxicity data on BBP to freshwater and marine fish are presented in Table 7.7. and the long-term toxicity data on BBP to fresh-water and marine fish are presented in Table 7.8.

Table 7.7
Short-term toxicity of Bytylbenzyl phthalate to fish

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

Table 7.8
Long-term toxicity of Bytylbenzyl phthalate to fish

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

In all tests performed, except for one acute test with Lepomis macrochirus, the effect concentrations of BBP were at concentrations equal to or below the water solubility of the substance. Furthermore, most of the effect concentrations determined are based on measured exposure concentrations. From the results obtained by Ozretich et al. (1983) /115/ with Cymatogaster aggregata, it can be seen that no further toxicity is obtained when the exposure period is increased from 96 h to 165 h. This indicates that steady state conditions and thus the maximum toxicity of BBP are reached during the 96 h test period.

7.3.5 Estrogenic effects

Jobling et al. (1995) /100/ studied the estrogenic effects of a range of chemicals, including BBP, 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. (1995) /100/ documented that BBP binds to the receptor, inhibiting the binding of natural estradiol. It has also been shown that BBP has mitogenic effect on the in vitro growth of human breast cancer cell (ZR-75) at test concentrations of 3.12 mg/l. In transiently transfected MCF 7 breast cancer cells, BBP was reported to affect the transcriptional activity of the estrogen receptor. BBP concentrations in the range from 0.31 to 31.2 mg/l stimulated the activity.

In a study by Sharpe et al. (1995) /102/, BBP was found to have estrogenic activity using a recombinant yeast screen. The relative potency of BBP was approx. 1·10⁶ times less than 17b-estradiol. In addition, it was found that BBP at a concentration of 10^-11 M increases the transcriptional activity in the presence of natural 17b-estradiol.

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. BBP was added to the drinking water of the pregnant female rats at low concentrations (1 mg/l). In adult life, males exposed in this way had testes that were reduced in size by 5-13% and a 10-21% reduction in their sperm production capacity. These effects were manifest in animals showing no gross changes. Meek et al. (1997) /103/ measured changes in reproductive organs of female rats. They showed that BBP produces no estrogenic activity. However, it has to be noted that the conclusions made by Meek et al. (1997) /103/ are built on unpublished data.

7.3.6 Summary and conclusions

Variable results were found in the toxicity tests with micro-organisms.

BBP has been shown acutely toxic (EC₅₀ or LC₅₀ values) to algae, crustaceans and fish in the range from 0.1 to 2.1 mg/l, in which algae seem slightly more sensitive than crustaceans and fish. BBP is thus considered very toxic to aquatic organisms.

NOEC levels in chronic toxicity tests with algae, crustaceans and fish were observed in the range from 0.03 to 0.35 mg/l.

BBP has shown to be estrogenic in vitro, stimulating human breast cancer cell growth and transcriptional activity of the estrogen receptor. BBP has been shown to significantly reduce testis size and sperm producing capacity of male rats exposed to low concentrations of the chemical during gestational and the lactational period. BBP has shown to be testis toxic in adult rats, causing atrophy of the testes, prostate, seminal vesicles and epididymis.

Lack of in vivo estrogenic effects studies in wildlife makes assessment of the potential estrogenic effects of BBP in wildlife difficult. However, data on male rats exposed to low concentrations of BBP during gestation and the lactation period indicate that estrogenic substances may affect the reproductive ability of adult male rats.

7.4 Environmental hazard classification

BBP has been shown acutely toxic (EC₅₀ or LC₅₀ values) to algae, crustaceans and fish in the range from 0.1 to 2.1 mg/l, in which algae seem slightly more sensitive than crustaceans and fish.

NOEC levels in chronic toxicity tests with algae, crustaceans and fish were observed in the range from 0.03 to 0.35 mg/l.

BBP is readily biodegradable in standard laboratory tests.

BBP is bioaccumulative in aquatic biota, which is demonstrated by the experimentally derived BCF value of up to 663 for fish.

The water solubility of BBP is » 2.7 mg/l which is 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 Butylbenzyl phthalate, it is proposed that BBP is classified "N; R50/53: Very toxic to aquatic organisms, may cause long-term adverse effects in the aquatic environment".

7.5 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 0.1 mg/l to a few mg/l. Long-term toxicity data are available on algae, crustaceans and fish with NOEC values in the range from 0.03 to 0.35 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 = 0.0001 mg/l.
 

8 Diisononyl Phthalate (DINP)

8.1 Physico-chemical properties

DINP (C₂₆H₄₂O⁴), CAS No.: 28553-12-0 and 68515-48-0, with an alkyl chain length of 9,9 /1/ is a colourless oily liquid. The molecular weight is 418.6 (418.6-432.6). DINP has a melting point of about -48°C and a boiling point at 440°C /1, 7/. The density is 0.97 g/ml and the vapour pressure is <5.0·10 mmHg at 25°C.

8.1.1 Water solubility

DINP is a high molecular weight phthalate. Evidence indicates that many of the measured water solubilities for high molecular weight phthalates esters reported in the literature are erroneously too high. The water solubility is calculated to be 2.31^-5 mg/l /7/ while, in the literature, several aqueous solubility data on DINP range from 7.8·10^-5 to 0.0006 mg/l. In a literature review by Staples et al. /1/, it was concluded that a water solubility of <0.001 mg/l was the most likely value based on available evidence.

8.1.2 Octanol-water partition coefficient

For high molecular weight phthalates as DINP, the HPLC method for determination of K_ow values cannot be used. Log K_ow values for DINP have thus been calculated by use of SPARC by USEPA /117/ and a value of 10.0 has been estimated while a value of 9.37 has been estimated in EPIWIN (1994) /7/. In a review by Staples et al. /1/, a log K_ow of >8.0 has been concluded as being the most likely value based on available evidence.

8.1.3 Summary

The approximate physico-chemical properties on DINP are summarized in Table 8.1.

Table 8.1
Physico-chemical properties of Diisononyl Phthalate (DINP)

8.2 Environmental concentrations and fate

8.2.1 Concentrations in the environment

The content of DINP in wastewater and sewage sludge from Danish treatment plants has been measured during recent years. An overview of the results is given in Table 8.2.

Table 8.2.
Diisononyl Phthalate (DINP) in treatment plants

It is the general picture that a high removal of DINP from the waste water is found during the wastewater treatment. However, mass balances show that most of the amount in the inlet water is found in sludge /9/, which roughly indicates that the degradation of DINP is limited and that DINP is adsorbed to and follows the sludge.

No data are available.

8.2.2 Abiotic degradation

No experimental data on the hydrolysis of DINP are available. The hydrolysis half-life at neutral pH and 25°C range is estimated to 3.4 years /7/.

No experimental data on photodegradation of DINP are available. Estimated photodegradation half-lives in the atmosphere are in the range from 0.2 to 2 days /4, 7/. In the aquatic environment, only insignificant photodegradation is expected /1/.

8.2.3 Biodegradation

Scholz et al. (1997) /65/ investigated the ready biodegradability of DINP in the Modified Sturm test (OECD 301B) and found a degradation of 79% after incubation for 28 days. Staples et al. (1996) /1/ refer to studies showing an ultimate biodegradation of <1-70% after incubation for 28 days.

When using an acclimated inoculum, Sugatt et al. (1984) /14/ demonstrated a biodegradability of DINP of 62% after 28 days.

Staples et al. (1996) /1/ refer to studies with DINP showing a primary biodegradability of more than 95% by employing a non-acclimated inoculum and 68->99% by using acclimated inocula.

Lundberg (1994) /118/ refers to a study for the Chemical Manufacturers Association on degradation of DINP. During a 3-week acclimation phase in a SCAS test, an average daily primary degradation of 68% was found and, in a succeeding die-away test, a primary degradability of more than 90% was found after 5 days of incubation.

No data are available.

Johnson et al. (1984) /68/ investigated the biodegradation of DINP in sediment which had been pre-exposed for 28 days before the start of the experiment. After 28 days of incubation, a primary degradation of 1% was determined under aerobic conditions and less than 1% under anaerobic conditions. Further experiments demonstrated that the same degree of degradation (1.2-1.6% after 28 days) was reached at various exposure concentrations from 0.02 to 10 mg/l. Finally, increasing degradation was observed with increasing temperature from about less than 1% at 12°C to more than 2% at 28°C.

8.2.4 Bioaccumulation

For DINP, only one bioaccumulation study performed with molluscs was found. A total BCF of 1844 was reported by Solbakken et al. (1985) /118/ (the exposure concentration was 61 mg/l and the test procedure was static).

8.2.5 Summary and conclusion

DINP seems to be efficiently removed from waste waters based on the low outlet concentrations compared to inlet concentrations in sewage treatment plants. However, mass balances show that most of the amount in the inlet water is found in sludge, which indicates that the degradation of DINP is limited and that DINP is adsorbed to and follows the sludge.

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

DINP exhibits a borderline ready biodegradability with some test results showing a mineralization greater than the pass level and some below the pass level. In a simulation test of a sewage treatment plant, a high primary biodegradability was found. This is inconsistent with the above conclusions based on measurements in full-scale wastewater treatment plants. In a sediment-water system, very low rates of primary biodegradation were found under both aerobic and anaerobic conditions.

DINP is bioaccumulative in aquatic biota, which is demonstrated by the experimentally derived BCF value of 1844 for molluscs.

8.3 Effects

8.3.1 Toxicity to micro-organisms

No toxicity with micro-organisms could be found.

8.3.2Toxicity to algae

One toxicity study with algae was found. The results are presented in Table 8.3.

Table 8.3
Toxicity of Diisononyl phthalate to algae

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

From the above results, DINP seems to have no acute or chronic toxicity to algae. The effect concentrations measured are, however, far above the water solubility of the substance.

8.3.3 Toxicity to invertebrates

The short-term toxicity data on DINP to freshwater and marine invertebrates are presented in Table 8.4 and the long-term toxicity data on DINP to freshwater and marine invertebrates are presented in Table 8.5.

Table 8.4
Short-term toxicity of Diisononyl phthalate to aquatic invertebrates

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

Table 8.5
Long-term toxicity of Diisononyl phthalate to aquatic invertebrates

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

From the above results, DINP seems to have no acute toxicity to crustaceans while a slight toxicity was observed in a 21 d survival test with Daphnia magna. Due to the low solubility of the substance, the effect observed may in part be ascribed to an indirect effect such as floating (entrapment) of the test animals or microdroplets which may adhere to the surface of the animals. The effect concentration measured is far above the water solubility of the substance.

8.3.4 Toxicity to fish

Only short-term toxicity data exist on fish. The toxicity data on DINP to freshwater and marine fish are presented in Table 8.6.

Table 8.6
Short-term toxicity data on Diisononyl phthalate to fish

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

DINP showed no acute toxicity in any of the acute toxicity tests performed. The NOEC values given are all far above the water solubility of the substance.

8.3.5 Estrogenic effects

In an investigation by Harris et al. (1997) /62/, DINP was shown to have weak estrogenic activity in in vitro recombinant yeast screen test, with a relative potency of approx. 5·10⁷ times less than 17b-estradiol.

8.3.6 Summary and conclusions

The results from ecotoxicology tests vary with a factor of up to 3. The reason for the variability should most probably be sought in experimental difficulties arising from the low water solubility of DINP. The formation of microdroplets, surface films and adsorption to surfaces of the test organisms lead to difficulties in maintaining steady exposure concentrations and/or cause direct physical effects.

DINP shows no acute toxicity to either algae, crustaceans or fish. Toxicity was observed in a long-term test with Daphnia magna (LOEC = 0.089 mg/l, NOEC = 0.034 mg/l). However, the toxicity observed is expected to be ascribed mainly to an indirect effect such as floating (entrapment) or microdroplets which may adhere to the surface of the animals.

Although no acute toxicity was observed at or above the water solubility of the substance in any of the tests performed, it has to be noted that the maximum toxicity of a substance will only be seen when steady state conditions have been achieved during the exposure time. With a log K_ow value of approx. 8, it can be estimated that steady state conditions in a fish test will not be reached before 1300 days of exposure /121/. However, taking the relatively high biotransformation of phthalatic esters into account, steady state conditions will probably never be reached for which reason maximum toxicity of DINP will never be seen.

As no chronic toxicity tests with fish are available, it cannot be excluded that DINP may cause long-term adverse effects at or below the water solubility of the substance in aquatic organisms (caused by direct or indirect effects).

In an investigation by Harris et al. (1997) /62/, DINP was shown to have weak estrogenic activity in in vitro recombinant yeast screen test, with a relative potency of approx. 5·10⁷ times less than 17b-estradiol.

No acute toxicity is observed at or below the water solubility level of DINP (» <0.001 mg/l).

Toxicity of DINP was observed in a long-term test with Daphnia magna (NOEC = 0.034 mg/l). However, the toxicity observed is expected to be ascribed mainly to an indirect effect such as floating (entrapment) or microdroplets which may adhere to the surface of the animals. No chronic or long-term tests performed with fish were available.

DINP is readily biodegradable in laboratory tests.

DINP is bioaccumulative in aquatic biota, which is demonstrated by the experimentally derived BCF value of 1844 for fish.

The water solubility of DINP is < 0.001 mg/l, which is well below 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 Diisononyl phthalate, it is proposed that DINP should not be classified as dangerous to the aquatic environment. However, it has to be noted that DINP is highly bioaccumulative.

8.5 PNEC for the aquatic compartment

No acute toxicity has been measured at concentrations at or below the solubility limit. Weak toxic effects were found in one test at a concentration far above the solubility limit. The available data do not allow a derivation of a PNEC_aquatic.
 

9 Diisodecyl Phthalate (DIDP)

9.1 Physico-chemical properties

DIDP (C₂₈H₄₆O₄), CAS No.: 26761-40-0 and 68515-49-1. DIDP is not a pure compound but a mixture of phthalates with side chains of average length 10,10 /6/. The molecular weight is about 446.7 (432.7-446.7) g. DIDP has a melting point of about -46°C and a boiling point at 463°C /1, 7/. The density is 0.961 g/ml and the vapour pressure is <5.0·10^-7 mmHg at 25°C /1/.

9.1.1 Water solubility

DIDP is a high molecular weight phthalate. Evidence indicates that many of the measured water solubilities for high molecular weight phthalate esters reported in literature are erroneously too high. In the literature, several aqueous solubility data on DIDP range from 7.4·10^-6 to <0.00013 mg/l. The solubility has been calculated to be 2.24·10^-6 /7/. In a literature review by Staples et al. /1/, it was concluded that a water solubility of <0.001 mg/l was the most likely value based on available evidence.

9.1.2 Octanol-water partition coefficient

For high molecular weight phthalates as DIDP, the HPLC method for determination of K_ow values cannot be used. Log K_ow values for DIDP have thus been calculated by use of SPARC by USEPA /117/ and a value of 10.0 has been estimated while a value of 10.36 has been calculated in /7/. In a review by Staples et al. /1/, a log K_ow of >8.0 has, however, been concluded as being the most likely value based on available evidence.

9.1.3 Summary

The approximate physico-chemical properties on DIDP are summarized in Table 9.1.

Table 9.1
Physico-chemical properties of Diisodectyl Phthalate (DIDP)

9.2 Environmental concentrations and fate

9.2.1 Concentrations in the environment

No data are available.

9.2.2 Abiotic degradation

No experimental data on the hydrolysis of DIDP are available. The hydrolysis half-life at neutral pH and 25°C range is estimated to 3.4 years /7/.

No experimental data on photodegradation of DIDP are available. The photodegradation half-life in the atmosphere is estimated to 0.2 days /7/. In the aquatic environment only insignificant photodegradation is expected /1/.

9.2.3 Biodegradation

The ready biodegradability of DIDP was determined in the OECD 301C test resulting in a degradation of 42% after 14 days of incubation /12/. Lundberg (1994) /118/ refers to a degradation of 30-100% of DIDP in the OECD 301C after 14 days of incubation. Staples et al. (1996) /1/ refers to two studies showing an ultimate biodegradation of 67% after 28 days of incubation.

When using an acclimated inoculum, Sugatt et al. (1984) /14/ demonstrated a biodegradability of DIDP of 56% after 28 days of exposure.

Staples et al. (1996) /1/ refer to studies with DIDP showing a primary biodegradability of 42% by employing a non-acclimated inoculum and 68->99% by using acclimated inocula.

Lundberg (1994) /118/ refers to a study for the Chemical Manufacturers Association on degradation of DIDP. During a 3-week acclimation phase in a SCAS test, an average daily primary degradation of 68% was found and, in a succeding die-away test, a primary degradability of more than 90% was found after 9 days of incubation.

No data are available.

In a sediment-water system, Johnson et al. (1984 /68/, cited by Lundberg 1994 /118/) found a mineralization of 1% after 28 days of incubation at 22°C.

9.2.4 Bioaccumulation

In the review given by Staples et al. (1996) /1/, several bioaccumulation studies with molluscs, crustaceans and fish are reported. Below, a number of studies on bioaccumulation of DIDP referred to in Staples et al. (1996) /1/ are given.

Brown & Thompson (1982a) /122/ determined the total BCF for Mytilus edulis, BCF values of 3977 and 2998 were found (exposure concentration: 4.4 and 41.7 µg/l, respectively; test procedure: flow through; exposure period: not known).

For crustaceans, only data on Daphnia magna exist. Brown & Thompson (1982b) /123/ determined the total BCF in static renewal tests and found the following BCF values: 90 (exposure concentration: 100.4 µg/l); 128 (exposure concentration: 32.6 µg/l); and 147 (exposure concentration: 9.6 µg/l).

Japan CITI (1992) /124/ found a total BCF for carp (Cyprinus carpio) of <3.6 to <14.4 (test procedure: flow through; exposure concentration: 0.1-1.0 µg/l; exposure period: not known).

9.2.5 Summary and conclusion

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

DINP exhibits a borderline ready biodegradability with some test results showing a mineralization greater than the pass level and some below the pass level. In simulation of a sewage treatment plant, a high primary biodegradability was found. No data are available on degradation under anaerobic conditions. In a sediment-water system, a very low primary biodegradation was seen under aerobic conditions.

DIDP is bioaccumulative in aquatic biota, which is demonstrated by the experimentally derived BCF value of up to 4000 for molluscs.

9.3.1 Toxicity to micro-organisms

Only one study with micro-organisms (protozoa) was found. The results of the test are given in Table 9.2.

Table 9.2
Toxicity of Diisodecyl phthalate to micro-organisms

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

No toxicity of DIDP to protozoans was found. The NOEC value given is far above the water solubility of the substance.

9.3.2 Toxicity to algae

One toxicity study with algae was found. The results are presented in Table 9.3.

Table 9.3
Toxicity of Diisodecyl phthalate to algae

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

From the above results, DIDP seems to have no acute or chronic toxicity to algae. The effect concentration measured is far above the water solubility of the substance.

9.3.3 Toxicity to invertebrates

The short-term toxicity data on DIDP to freshwater and marine invertebrates are presented in Table 9.4 and the long-term toxicity data on DIDP to freshwater and marine invertebrates are presented in Table 9.5.

Table 9.4
Short-term toxicity of Diisodecyl phthalate to aquatic invertebrates

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

Table 9.5
Long-term toxicity of Diisodecyl phthalate to aquatic invertebrates

F: Fresh water; S: Salt water; N: Nominal; M: Measured concentration; NOEC: No Observed Effect Concentration; LOEC: Lowest Observed Effect Concentration.

From the above results, DIDP seems to have no acute toxicity to crustaceans while a slight toxicity was observed in a 21 d survival test with Daphnia magna. Due to the low solubility of the substance, the effect observed may in part be ascribed to an indirect effect such as floating (entrapment) or microdroplets which may adhere to the surface of the animals. The effect concentration measured is far above the water solubility of the substance.

9.3.4 Toxicity to fish

Only short-term toxicity data exist on fish. The toxicity data on DIDP to freshwater and marine fish are presented in Table 9.6.

Table 9.6
Short-term toxicity data on Diisodecyl phthalate to fish