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The environmental fate of phthalate esters: A literature review
Chemosphere, Vol. 35, No. 4, pp. 667-749, 1997
Pergamon PII: S0045-6535(97)00195-1
0 1997 Elsevier Science Ltd All rights reserved. Printed in Great Britain 0045-6535/97 $17.00+0.00
THE ENVIRONMENTAL FATE OF PHTHALATE ESTERS: A LITERATURE REVIEW
Charles A. Staples ~, Dennis R. Peterson2, Thomas F. Parkerton2 and William J. Adams3
Assessment Technologies, Inc., 10201 Lee Highway, Suite 305, Falrfax, VA 22030 USA I Exxon Biomedical Sciences, Inc., Mettlers Road, CN 2350, East Millstone, NJ 08875 USA2 Kennecott Utah Copper Corp., Magna, UT 84044 USA 3 (Received in Germany 4 February 1997; accepted 5 March 1997)
ABSTRACT A comprehensive and critical review was performed on the environmentalfate of eighieen commercial phth~ate esters with alkyl chains ranging from 1 to 13 carbons. A synthesis of the extensive literature data on physicoehemical properties, partitioning behavior, abiotic and biotic transformations and bioaccumulationprocesses of these chemicals is presented. This chemical class exhibits an eight order of magnitude increase in cctanol-water partition coefficients (K~) and a four order of magnitude decrease in vapor pressure (VP) as alkyl chain length increases from 1 to 13 carbons. A critical review of water solubility measurements for higher molecular weight phtbadate esters (i.e. alkyl chains ~ 6 carbons) reveals that most published values exceed true water solubilities due to experimental difficulties associated with solubility determinations for these hydrophobic organic liquids. Laboratory and field studies show that partitioning to suspended solids, soils, sediments and aerosols increase as K ~ increases and VP decreases. Photodegradation via free radical attack is expected to be the dominant degradation pathway in the atmosphere with predicted half-lives of ca. 1 day for most of the phtl~ate esters investigated. Numerous studies indicate that phfludate esters are degraded by a wide range of bacteria and aetinomyeetes under both aerobic and anaerobic conditions. Standardized aerobic biodegradation tests with sewage sludge inocula show that phthalate esters undergo ~ 50% ultimate degradation within 28 days. Biodegradation is expected tobe the dominant loss mechanism in surface waters, soils and sediments. Primary degradation half-lives in surface and marine waters range from <1 day to 2 weeks and in soils from <1 week to several months. Longer half-lives may occur in anaerobic, oligotrophie, or cold environments. Numerous experiments have shown that the bioaccnmulationof phthalate esters in the aquatic and terrestrial foodchaln is limited by biotransformation, which increases with increasing trophic level. Consequently, models that ignore biotransformation grossly exaggerate bioaccumulation potential of higher molecular weight phthalate esters. This review provides the logical first step in elucidating multimedia exposure to phthalate esters. © 1997 Elsevier Science Ltd 667
668 INTRODUCTION ~01amaercial Uses and Products Phthalate esters are widely used industrial chemicals (ECETOC, 1988; Giam et al., 1984; Graham, 1973), serving as important additives which impart flexibility in polyvinylchioride (PVC) resins. Phthalate esters are also used to varying degrees in other resins such as polyvinyl acetates, cellulosics, and polyurethanes. The stability, fluidity and low volatility of higher molecular weight phthalate esters make them highly suitable as plasticizers.
Most of the mid- to high- molecular weight phthalate esters are used in the manufacturing of PVC, while di-n-butyl phthalate is used in epoxy resins and cellulose esters and specialized adhesive formulations. Dimethyl and diethyl phthalate esters are typically used in cellulose ester-based plastics, such as cellulose acetate and butyrate. Plasticizers are used in building materials, home furnishings, transportation, clothing, and to a limited extent in food (packaging) and medical products. Release of phthalate esters into the environment during manufacture, use, and disposal has recently been reviewed (Warns, 1987; Cadogan et al. 1993). The intent of this review is to evaluate the extensive literature on the physicochemical properties and the abiotic/biotic processes that dictate the ultimate fate and distribution of phthalate esters in the environment.
The eighteen commercially important phthalate esters reviewed are listed in Table 1, where the abbreviations used for the various phthalate esters are presented. In general, the abbreviation indicates when the phthalate ester is a mixture of branched or linear isomers (i.e., DnOP for linear di-n-octyl phthalate and DIOP for isomeric diisooctyl phthalate). Butylbenzyl phthalate (BBP), is non-symmetric, having both butyl and benzyl alcohol esters groups. With the exception of di-2(ethylhexyl) phthalate, which is a pure isomer, higher molecular weight phthalates (i.e. alkyl chains >6 carbons) are mixtures based on the alcohols used for synthesis. For example, 711P is a mixture of di-heptyl, nonyl, and andecyl phthalates. Likewise, 610P is a mixture of hexyl, octyl, and decyl isomers. Other products, such as diisononyl and diisodecyl phthalates are isomeric mixtures comprised principally of C9 or C10 alkyl chains. Considerable differences in structure are thus possible for certain commercial phthalate esters.
PHYSICOCHEMICAL PROPERTIES
Review of phthalate ester physical properties is important to help understand behavior and fate in the environment. A number of the important physicochemical properties of phthalate esters are reviewed below.
Common commercial phthalate esters are liquids at ambient temperatures (Table 2). Nearly all have melting points (or pour points) below -25 °C. The exceptions are DMP, DUP, and 610P with melting points of +5.5 °C, -9 °C, and -4 ° C,
669 Table 1. Eighteen Commercial Phthalate Esters Reviewed. Abbreviation
Phthalate Ester
CAS No.
DMP
Dimethyl Phthalate
131-11-3
DEP
Diethyl Phthalate
84-66-2
DAP
Diallyl Phthalate
131-17-9
DPP
Dipropyl Phthalate
131-16-8
DnBP
Di-n-Butyl Phthalate
84-74-2
DIBP
Diisobutyl Phthalate
84-69-5
BBP
Butylbenzyl Phthalate
85-68-7
DHP
Dihexyl Phthalate
84-75-3; 68515-50-4
DnOP
Di-n-Octyl Phthalate
117-84-0
BOP
Butyl 2-Ethylhexyl Phthalate
85-69-8
610P
Di (n-Hexyl, n-Octyl, n-Decyl) Phthalate
25724-58-7; 68515-5 I-5
DEHP
Di (2-Ethylhexyl) Phthalate
117-81-7
DIOP
Diisooctyl Phthalate
27554-26-3
DINP
Diisononyl Phthalate
28553-12-0; 68515-48-0
DIDP
Diisodecyl Phthalate
26761-40-0; 68515-49-1
D711P
Di (Heptyl, Nonyl, Undecyl) Phthalate
3648-20-2; 68515-44-6 68515-45-7; 111381-89-6 111381-90-9; 111381-91-0
DUP
Diundecyl Phthalate
3648-20-2
DTDP
Ditridecyl Phthalate
119-06-2; 68515-47-9
respectively. Phthalate esters have boiling points varying from about 230°C to 486°C. Higher molecular weight phthalate esters have boiling points that must be determined at reduced pressure to prevent thermal decomposition. The low melting point and high boiling point of these phthalate esters contribute to their usefulness as plasticisers, heat transfer fluids, and carders.
Water Soinbilitv Water solubility is an extremely important property that influences the biodegradation and bioaceumulation potential of a chemical, as well as aquatic toxicity. Water solubility also is a determining factor controlling the environmental distribution of chemicals. Losses from wastewater treatment facilities, landfills and sludge-amended soils are partially a function of aqueous solubility.
4 4, 6 ~
CI2HI404
CI4HI404
Ci4His04
Ct6H2204
C16H2204
DEP
DAP
DPP
DnBP
DIBP
7, 9,11 11
C24H3804
C~H3004
CzsH4004
C24H3sO4
C24H3sO4
C26H~204
C28H4604
C26H4204
C30H~O4
C34H5804
DnOP
BOP
610P
DEHP
DIOP
DINP
DIDP
D711P
DUP
DTDP
530.8 {506.8-544.8
447.7 {432.7-474.7
418.6 {362.6-474.7
446.7 {432.7-446.7
418.6 {418.6-432.6
390.6 {376.6-390.6
390.6
404.6 {334-447}
334.4 {278.4-390.6
390.6
334.4
312.4
278.4
278.4
250.3
246.2
222.2
194.2
Molecular Weight
1.111 1.011 0.978
-35 -27.4 -25
-37
-9
<-50
-46
-48
-46
-47
-4
0.953
0.96
0.97
0.961
0.97
0.986
0.986
0.97
--
1.050 c
-58
-37
1.042
1.118
-40
-35
1.192
Specific Gravity (20 °C)
5.5
Melting Point (°C)
Data from Howard et al. (1985); ttoward et al. (I 989); Sears and Touchette (1982); and manufacturer's unpublished data. For mixtures, the formula for the most representative isomer is provided. Aromatic ring { } molecular weight range of isomer At 15 °C
13
10
9
8
8
6, 8,10
6 c, 8
8
6
CI9H1oO4
C~H3oO4
BBP
DHP
4
3
3
2
1
CioHioO4
DMP
Alkyl Chain Length
Formula b
Compound
Table 2. Physical Properties of Eighteen Phthalate Esters?
.,.....I O
671 Precise water solubility measurements for chemicals with moderate to high water solubilities (> I mg/L) can be obtained easily with conventional methods e.g., shake-flask procedures. In contrast, water solubility measurements for more hydrophobie compounds can be confounded by a variety of experiment problems. First, the inability to separate colloidal emulsions of undissolved chemical from the aqueous phase may cause difficulties in determining accurate water solubilities (Yalkowsky and Banerjee, 1992). Another source of error of particular concern for phthalate esters is artifactual contaminationfrom laboratory plastics (Giam et al., 1984). Since higher molecular weight phthalate esters are slightly less dense liquids than water at ambient temperatures, these chemicals form surface films at the air-water interface. Consequently, errors may be introduced when withdrawing aqueous phase samples for analysis through surface films (Howard et al., 1985). Each of the above problems can lead to experimental artifacts that yield measured values that overestimate the true water solubility.
Recognizing the above limitations, reported aqueous solubilities ofphthalate esters are summarized in Table 3. Results show a declining trend in water solubility with increased carbon number of the alcohol moiety for the lower molecular weight phthalates through DHP. Moreover, independent experimental measurements are generally in good agreement and are consistent with theoretical predictions obtained by the SPARC (Karickhoffet al., 1991) as well as the EPIWIN (Meylan and Howard, 1995) structure-activity models. Thus, experimental measurements are believed to be reliable for lower molecular weight phthalates.
In contrast, experimental data for higher molecular weight phthalates often appear suspect. First, the expected decreasing trend in water solubility with increasing alkyl chain length is not apparent for these compounds (Howard et al., 1985). Furthermore, measurements obtained by independent investigators using different methods often vary by several orders of magnitude. For example, Howard et al. (1985) reported a value of 0.34 mg/L for DEHP using a shake-flask/centrifugation method. Similar results of 0.285 mg/L and 0.36 mg/L were reported by HoUifield (1979) and Defoe et al. (1990) respectively, using uephelometric techniques. In comparisojj, Leyder and Boulanger (1983) measured a water solubility for DEHP of 0.041 mg/L using a shake flask/centrifugation method. In seawater, Giam et al. (1980) and Howard et al. (1985) report DEHP solubilities of 1.16 and 0.160 rag/L, respectively, using conventional techniques. In contrast, Boese (1984) reported a value of 0.0006 mg/L in seawater using a generator column. This latter value is consistent with the predicted estimate of 0.0026 mg/L and 0.0011 mg/L obtained from the SPARC and EPIWIN models, respectively (Long, 1995; Meylan and Howard, 1995). Early studies reported that the water solubility of DINP and DIDP ranged from 0.2 - 1.19 mg/L. More recent experiments that employed a slow-stir shake flask method and included precautions to prevent sampling of ulidissolved test material indicate values of 0.0006 and <0.00013 for DINP and DIDP, respectively (Exxon Biomedical Sciences, Inc., 1996a). These measurements are in much closer agreement to theoretical predictions based on structure activity relationships (Table 3).
672 T a b l e 3. S e l e c t e d P a r a m e t e r s C o n t r o l l i n g t h e E n v i r o n m e n t a l Compound
............
DMP
1.46 (1) 1.47 (2) 1.48 b 1.53 (3) 1.56 (4) 1.60 (37) 1.61 (4)
Log K . - . . . . . . . . . . . . . . . . . . . . . . 1.61 (5) 1.62 (6) 1.62 (7) 1.66 (5) 1.66 d 1.74 (5) 1.90 (2)
2179 d 2810 (1) 3160 (2)' 3300 b 4000 (2)
D i s t r i b u t i o n o f P h t h a l a t e E s t e r s a.
AQ SOL (mg/L) . . . . . . . . . . . . . . . . . 4248 (4) 4290 (3) 4320 (18) 4200*
V P (mm Hg, ~25 °C) . . . .
1.65E-3 (2) 1.8E-3 (27) 5.4E-3 (38) 5.5E-3 c 2.9E-2 d 2.0E-3"
1.61' DEP
2.21 2.21 2.24 2.29 2.35
(6) (6) (2) (2) (3)
2.42 (37) 2.47 (8) 2.51 b 2.65 a 2.67 (4) 3.00 (9)
260 d 400 ~ 680 (19) 720 (20) 896 (18) 928 (3)
1080 (2) 4.8E-5 (27) 1160-1240 (21) 20°C 3.9E-4 (28) 1340-1400 (21) 30°C 6.1E-4 (28) 7028 (4) 8.1E-4' 1.2E-3 (35) 1100" 1.65E-3 (2) 5.0E-3 a
43 d
182 (3)*
3.7E-5 ¢ 1.6E-4 (35)*
108 (3)*
8.9E-5 c 1.04E-3 d
2.38* 1.0E-3* DAP
3.16 b 3.23 (3)*
3.36 d
DPP
3.27 (3)* 3.57 b
3.63 d
38 a 47 ~
DnBP
3.74 (2) 4.08 (10) 4,11 (11) 4.13 (6) 4.30 (12) 4.39 (10) 4.50 (37) 4.56 (10)
4,57 (3) 4.61 d 4.63 b 4.72 (8) 4.72 (13) 4.79 (2) 5.15 (31)
1.Y 11.2 (2) 3.2 (26) 13.0 (18) 3.25 (22)" 4.9 b 11.2' 8.%9.6 (13) 10.l (3) 10.5- I 1.1 (21 ) 20 °C 11.2-11.5 (21) 30°C
9.3E-6 (35) 1.7E-5 (32) 1.9E-5 (29) 2.08E-5 (30) 2.7E-5 (27) 3.5E-5 (34)
4.46 e
5.1 d 6.2 (23) 9,6 b
20.0(24) 20.3 (3)
1.8E-6 c 5.8E.-4 e
4.77 (11) 4.77 (14) 4.77 b 4.84 d 4.91 (3)
0.67 e 0.70 (23) 1.90 (20) 2.40 b 2.69 (2)
2.82 2.90 2.90 40.2
100b
3.6E-5 (36)
4.45* DIBP
4.11 (3) 4.31 b 4. I l *
BBP
3.57 3.97 4.05 4.11 4.73 4,75
(2) (6) (4) (11 ) (37) (4)
20.0 *
(3) (18) (5) (4)
2.7* 4.59*
DHP
8.7E-7 (35) 7,7E-6 e 8,6E-6 (14) 9.0E-6 (2) 9.1E-5' 5.0E-6"
5.65 (2)
6.67 ~
0.019 d
1.9E-8 ¢
5.93 (2) 6.57 e
6.82 (27)
0.049 b 0.24 (2)
1.8E-6 (34) 1.4E-5 (2) 1.2E-5 d
6.30* 0.05*
5.0E-6"
1.16E-3 d
4.1E-5 (30) 7.3E-5 (2) 1.2E-4 c 2.5E-4 d 2.7E-5"
673 Table 3. Selected Parameters Controlling the Environmental Distribution o f Phthaiate Esters (cont.) Compound DnOP
.........
Log K ,
5,22 (3) 7,06 (15) 8.06 (6)
8.10 (37) 8.30b ~,~4 d
AQ S O L (mg/L) . . . . . . .
VP (ram Hg) - ~
0.00046b 0.0005~ 0.001 (20)
0.02 (13) 0.02 (23) 3.00 (181 0.0005*
2.2E-7 (27) <3.4E-7 (35) 3.4E-7d
0.02d 0.11b*
1.1E-7 (35)* 2.4E-5d
1.2E-4~
8.06*
1.3E-5¢ 1.9E-4 (29) 1.0E-7*
BOP
6.28b* 6.5d
610P
5.9-8.6 (2) 7.25*
8.54d
0.0004d 0.05*
0.90 (2)
3.4E=7d 4.9E-6 (2)*
4.9E-4¢
DEHP
4.20-5.22 (16) 5.11 (ll) 7.06 (151 7.14±0,15 (16) 7.27 (37)
7.54b 7.80-8.90 (17) 7.86-8,15 (16) 7.94 (2) 8.06 (6)
0.0006 (25)' 0.0011 d 0.0026 b 0.041 (3) 0.16 (2)®
0.285 (23) 0.34(2) 0.40 (17) 0.40 (18) 1.16 (22)'
4.1E-8 4.5E-8 7.2E-8 9.8E-8 2.8E-7
<3.4E-7 (34) 3.8E-7 (27) 7.1E-7 (32) 6.4E-6 (2) 1.4E-6d
7.45 (15)
~,~9 d
0.27-0.36 (13)
1.2 (26")°
3.3E-7 (36)
7.45±0.06 (16)
7.50*
DIOP
8.00b* 8.39~
DINP
>8.0 (2)* 9.0b
0.003* 0.00024d 0.00081 b
0.09 (2)
1.4E-4~ 1.0E-7"
2.0E- I0c 3.4E-7 (35) 1.4E-6d
1.4E-6 (29) 5.6E-6 (21 1.0E-6*
0.0006 (39) 0.20 (21 <0.001 *
5.4E-7 (2)
2.2E-6 d 7.4E-6 ~
0.28 (24) 1.19 f21
5.1E-8 (36) 5.6E-8 (35)
0.001" 9.4d
(33) (12) (35) (I I) (29)
7.8E-5 b 2.3E-5 ~
<5.0E-7 *
DIDP
>8.0 (2)* I0,0b
<0.00013 (39)
<0.001*
<5.0E-7 (2)
<5,0E-7"
71 lP
6.0 (2) >8.0 (2)*
9.52d
1.7E-5d <1.0 (2)
<0.001"
5.0E-8d 2.8E-7 ( 3 5 )
<5.0E-7 (2"J <5.0E-7'
DUP
>8.0 (2)* 11.2b
11.5d
1.6E-7t' 4.2E-7b
1.10 (2)
5.3E-7 (35)
DTDP
>8.0 (2)* 13.1b
13.4d
1.5E-9d 4.2E-9b <0.3 (2)
10.3d
<0.001'
b ' d ° *=
0.34 (23) <0.001*
2.5E-11d
3.7E-8d
<5.0E-7' <5.0E-7 (2) <5.0E-7"
Includes measured and calculated values. Values calculated from the SPARC Structure Activity Program (Long, 19951. Calculated using the structure activity relationships in A S T E R (Russom et al., 1991). Calculated using the structure activity relationships in EPIWIN (Meylan and Howard, 1995). Solubility measurements made in seawater. R e c o m m e n d e d values based on available evidence.
References 1. Nielsen and Bundgaard, 1989; 2. Howard et al. 1985; 3. Leyder and Boulanger, 1983; 4. Veith et al. 1989; 5. Renberg et al. 1985; 6. McDuffie, 1981; 7. _P~dsforth, 1986; 8. Hansch and Leo, 1987; 9. DeKoch and Lord, 1987; 10. Harnish et ul., 1983; 11. OECD, 1981; 12. Haky and Leja, 1986; 13. DeFoe et al. 1990; 14. Gledhill et ,I. 1980; 15. DeBruijn and Hermens, 1989; 16. Brooke et al. 1990; 17. Klein et al. 1988; 18. Wolfe et al. 1980a; 19. R u u e l l and McDuffie, 1986; 20. Tanii and Hashimoto, 1982; 21. Schwarz and Miller, 1980; 22. Giam et al. 1986; 23. Hollifield, 1979; 24. C M A , 1983; 25. Boese, 1984; 26. Sullivan et al. 1981; 27. Stephen.son and Malanowski, 1987; 28. Grayson and Fosbraey, 1982; 29. Wemer, 1952; 30. G-uckel et al. 1982; 31. Veith et al. 19791); 32. Dobbs et ai. 1984; 33. Dobbs and Cull, 1982; 34. Fdssel, 1956; 35. Sears and Darby, 1982; 36. Quackenbos, 1954; 37, Ellington and Floyd, in press; 38. Cowen and Baynes, 19g0; 39. Exxon Biomedical Sciences, Inc., 1996a.
674 Other evidence indicates many of the measured water solubilities for high molecular weight phthalate esters reported in the literature are erroneously too high. First, octanol-water partition coefficients obtained by the most reliable experimental techniques (i.e. generator column or "slow-stir" shake flask methods) indicate much higher values when compared to conventional methods. Numerous studies using reliable methods (as discussed in the next section) indicate that the log Ko,vof DEHP is consistently in the 7.0-7.8 range (De Bruijn and Hermens, 1989; Brooke et al., 1990; Ellington and Floyd, in press). However, early Kowmeasurements obtained using less reliable conventional shake flask procedures yielded values two orders of magnitude lower (OECD, 1981; Brooke et al., 1990). Given the well established inverse relationship between true water solubility and Kow, these results suggest that historical water solubility determinations obtained by conventional methods are likely subject to large systematic errors.
Lack of aquatic toxicity provides additional evidence supporting the view that the true water solubilities of higher molecular weight phthalate esters are lower than most data suggest (Adams et al., 1995, Rhodes et al., 1995; Staples et al., 1996). The non-toxic nature of higher molecular weight phthalate esters implies that the true water solubility must be sufficiently low to prevent test organisms from achieving a critical body burden that elicits adverse effects. Given the imprecision of experimental measurements, coupled with the inconsistencies between experimental data and 1) theoretical predictions based on structure-activity; 2) experimental Kow measurements obtained using the most reliable methods available; and 3) aquatic toxicity information, water solubility measurements for higher molecular weight phthalate esters (Table 3) should be viewed as suspect. Predictions based on structure-activity relationships and from slow-stir measurements of log Koware believed to provide more reliable estimates than most data obtained using conventional shake flask or HPLC methods. Further research to confirm the validity of these predictions is needed.
Octanol/Water Partitionin~ The equilibrium distribution of an organic chemical between water and octanol (K~) is an important physical constant for predicting the tendency of a chemical to partition to water, animal lipids, sediment, and soil organic matter. Numerous correlations describing relationships between K~wand soil sorption, water solubility, bioconcentration, and toxicity have been developed (Lyman, 1982; DeWolfe et al., 1992; Veith et al., 1980; Gossett et al., 1983; and Geyer et al., 1984).
An early method of determining K,w was to agitate test chemical in a two-phase mixture of octanol and water and measure the equilibrium concentration of the chemical in both phases (OECD, 1981). This simple design contains a number of technical difficulties. Equilibrium must be reached and adequate separation of phases attained. The equilibrium eoneenlration of the test chemical in both phases must be below its maximum solubility in each phase. For highly hydrophobic and low solubility chemicals such as high molecular weight phthalate esters, the concentration will be substantially higher in one phase making accurate measurements in both phases dittleult. The use of radionetive test
675 chemical avoids some of these practical difficulties but introduces new problems in that the radioactivity measured at low activities may be difficult to interpret due to traces of water soluble impurities. Chemical transformations may also confound results.
A number of predictive methods have been developed to estimate octanol-water partition coefficients (Veith et al., 1979b; Lyman et al., 1982; DeKoch and Lord, 1987; McDuffie, 1981; Renberg, et al., 1985; Klein et al., 1988; Doucette and Andren, 1988; and Long, 1995). Many of these methods relate K~,,to other physical properties. Another approach is to estimate K~v based on HPLC retention time. Howard et al. 0985) reported results for a number of lower molecular weight phthalate esters using the HPLC method; however, the HPLC method could not be used to determine log K~ for the higher molecular weight phthalate esters DIOP, DIDP, DUP, and DTDP because these phthalate esters did not elute from the column under test conditions. Thus, higher molecular weight phthalate esters are too hydrophobic to quantify. Ko,, by this method.
The simplest and currently most reliable method for determining K~s for low solubility, hydrophobic compounds is the "slow-stir" method (DeBruijn and Hermens, 1989; Brooke et al., 1990; and Ellington and Floyd, in press). This procedure allows equilibrium while minimizing the formation of emulsions that can confound exTer:rnen~l results.
A summary of published Kowmeasurements ofphthalate esters are presented in Table 3. Fairly cons~ste'nt results are seen for the lower molecular weight phthalate esters (DMP, DEP, DnBP, and BBP). With increasing alkyl chain length, the log K~ increases indicating greater hydrophobicity, as expected. The median log K~ values for DMP, DEP, DnBP, and BBP were 1.61,2.38, 4.45, and 4.59, respectively. DEHP is the most widely studied phthalate ester of the higher molecular weight phthalate esters. Howard et al. (1985) estimated a log K ~ for DEHP of 7.94 using HPLC. Measurements using the slow-stir technique ranged from 7.14 to 7.45 (Brooke et al., 1990; DeBruijn and Hermens, 1989; Ellington and Floyd, in press). The U.S. EPA estimated a log K ~ of 7.54 for DEI-IP based on molecular structure using the SPARC model (Long, 1995). Log I~,,, values for higher molecular weight phthalate esters DrNP through DTDP were estimated using SPARC to be even higher, ranging from 9 to 13 (Ellington and Floyd, in press; Long, 1995). Similar estimates were obtained using the EPIWIN Model (Meylan and Howard, 1995).
Vapor Pressure
Vapor pressure plays an important role in the fate of ~gifive emissions and other releases of ph~flate esters to the atmosphere. Vapor pressure is typically deW.tined by direct pressure m~surement at elevated temperatures. Such data may be extrapolated to estimate vapor pressure at ambient temperatures using either the Clausius-Clapeyron or Antoine equations. The direct pressure measurement technique is limited by the sensitivity of the pressure measuring
676 device for poorly volatile substances. Since it is non-selective, the vapor pressure of the most volatile component of mixtures is measured.
Another method for vapor pressure measurement of poorly volatile substances is the gas saturation technique (U.S. EPA, 1980). It relies on a saturator column to generate a continuous stream of vapor saturated air at ambient temperature. The vapor is trapped and accumulated on a sorbent which is ultimately desorbed and analyzed chromatographically. This assures that the method is substance specific and not invalidated by the presence of volatile contaminants. Howard et al. (1985) utilized the gas saturation method to measure the vapor pressure of a series of phthalate esters. However, since test conditions were at 25°C, rather than elevated temperatures, the sensitivity of the method may be limited by the low air concentrations of poorly volatile chemicals.
Vapor pressure measurements reported in the literature for phthalate esters are summarized in Table 3. Measured values obtained in different studies often vary by more than one order of magnitude. However, a general trend is apparent showing that the vapor pressure of phthalate esters decline more than four orders of magnitude with increasing alkyl chain length. Comparison of measured data with the model predictions obtained by structure activity relationships included in the ASTER (Russom et al., 1991) and EPIWlN (Meylan and Howard, 1995) models often indicate more than order of magnitude discrepancies (Table 3).
ENVIRONMENTAL PARTITIONING Air-Water Partitioninc, A chemical's equilibrium distribution between water and air serves as a guide to estimate the tendency of a substance to escape from water into air. The ratio of the vapor pressure to the molar water solubility estimates the Henry's Law constant, which is the measure of the equilibrium distribution coefficient (Thomas, 1982). Henry's Law constants (H) were calculated from the recommended vapor pressure and water solubility data shown in Table 3.
For lower molecular weight phthalate esters (DIvIF, DEP, DAF, DPP, DnBP, DIBP, and BBP), H values ranged from 1.2E=7 to 8.8E-7 atm-m3/mole. Compounds with H values in the range of 1.0E=? a~n=m3/mole are generally considered to have negligible volatility (Howard et al., 1985). For all higher molecular weight phthalate esters except BOP, calculated H values ranged from about 1.7E=5 to 5.5E-4 atm-m3/mole. The higher H values are due to the greater decrease in water solubility relative to vapor pressure with increasing alkyl chain length. The H values for some of the higher molecular weight phthalate esters (DINF, DIDP, 71 IP, DUP, and DTDP) are difficult to calculate due to the extremely low vapor pressures and water solubilities that are not accurately known.
677 Vopor-Aerosol Partitioninv The distribution of organic chemicals in air between gaseous and particulate phases can be estimated by the Junge model (Eisenreich et al., 1981): 0
-
cO Po*cO
m
(I)
where 00 (phi) is the fraction of chemical on the particulate (i.e. aerosol) phase, c is a constant (approximately 0.13 for many organic chemicals), 0 is the particle surface area per unit volume and Po is the liquid vapor pressure in mm Hg at the temperature of interest. Assuming a typical suspended particulate concentration in air of 20-40 ug/m3 and a surface area of 1-3 m2/g, 0 is approximately 3 x 10"6cm-'/cm3 (Strachan and Eisenreich, 1988). By substituting this value and the vapor pressures of the individual phthalate esters reported in Table 3 into equation (1), ~ values for each phthaiate ester can be determined. Results are expressed as a dimensionless decimal fraction between 0 and I. Therefore, as 00 decreases, more chemical resides in the vapor phase. Results indicate that phthalate esters with alkyl chain lengths of less than six carbons (DMP through BBP) exist primarily in the vapor state whereas compounds with longer alkyl chains are mainly associated with aerosols (Table 4). This simple calculation does not take into account a variety of factors that can affect partitioning such as variability in particulate concentrations or surface area or the influence of temperature on the vapor pressure. Nevertheless, this calculation illustrates trends in the relative importance of atmospheric partitioning for different phthalate esters.
Partitioningbehavior influences the removal of contaminants from airvia washout by rain or snow. The washout ratio (W) is defined as the dimensionless ratioof chemical concentrations in precipitationto that in air.Washout can occur by both vapor and particle scavenging mechanisms and can be expressed as (Biddleman, 1988):
w = (1-00) ~ + 00 w
RT = (1-oo) - - f f , 00
(2)
where W, is the vapor washout ratio, RT/H in the reciprocal of the dimensionless Henry's constant at the temperature of interest, W, is the particle scavenging coefficient and 00 was defined in equation (1). Based on measurements for selected phthalate esters reported by Ligocki et al. (1985), a representative value for Wpis 20,000. Using this estimate and Hand 00 values reported in Table 4, equation (2) was used to calculate washout ratios (W) for each phthalate ester (Table 4). Limited field data indicate washout ratios for DEP, DnBP and DEHP range from 1000 to 100,000 (Arias and Glare, 1981; Ligocki et al., 1985; Atl~ and Glare, 1988; Thuren and Larsson, 1990). Thus field measurements are of the same order of magnitude as simple model predictions obtained using equation (2). One key factor contributing to the variability of field-determined washout ratios appears to be the temperature dependence of the vapor pressure. During the wanner summer months, the vapor pressure increases, which in accordance to equation (1), results in
678 Table 4. Selected Parameters Determining the Environmental Distribution of Phthalate Esters. Compound Henry's Constant (atm-m3/mol)
Ko, (L/kg) (soil/sediment)
K,~ (L/kg) suspended solids
dp
Washout Ratio (W)
DMP
1.22E-07
55 (1) 8O-36O (2)
<50,000 (8)"
0.00019
2.0E+5
DEP
2.66E-07
69 (3) 704-1726 (4)
79,400 (8) a
0.00039
9.0E+4
DAP
2.85E-07
0.0024
8.6E+4
DPP
3.05E-07
0.0039
8.0E+4
DnBP
8.83E-07
0.014
2.8E+4
1020 (10) a
0.038
1.3E+5
100,000 (8)"
0.072
3.1E+4
0.072
2.0E+3
0.80
1.6E+4
0.80
2.8E+4
0.80
1.6E+4
DIBP
1.83E-07
BBP
7.61E-07
1375 (3) 14,900 (5)
17,000 (3)
158,500 (8) ~ 39,800 (9) 1230 (10) ~ 7200 (11)
9ooo (6) DHP
4.40E-05
DnOP
1.03E-04
BOP
4.00E-07
DEHP
1.71E-05
52,600 (7) 2.0E+6 (8)"
87,420 (3) 482,000 (7) 510,000 (5) •
DIDP
286,000 (7)
DTDP
1.2E+6 (7)
1.0E+6 (8) a 398,000 (9) 22,000 (1 O)a 583,000 (11)
Calculated from Kd assuming a 0.10 organic carbon fraction. 1. Osipoffet al. 1981; 2. Banerjee et al. 1985.3. Russell and McDuffie, 1986; 4. von Oepen et al. 1991; 5. Sullivan et al. 1982; 6. Gledhi!l et al. 1980; 7. Williams et al. 1995; 8, Furtmarm, 1993; 9. Germain and Langlois, 1988; 10, Preston and A1-Omran, 1989; 11. Ritsema et al., 1989.
reduced partitioning to aerosols and hence lower washout ratios. In contrast, during the colder winter months, the vapor pressure is reduced so that partitioning to aerosols is enhanced resulting in more efficient particle scavenging and thus higher washout ratios. This temperature-dependent phenomenon has been observed in Sweden for the atmospheric removal of both DnBP and DEHP (Thuren and Larsson, 1990).
Water-Solid Partitionin~ The sorption of phthalate esters to soil, sediment, or suspended solids is partially governed by the relative hydrophobicity of the chemical. Such hydrophobic chemicals adsorb principally to the organic matter associated with
679 the solid. Adsorption is generally measured by shake flask techniques in which a solid, water and test chemical are agitated in a closed system. Complexities arise in interpreting solid sorption test results. Sorption to soil, sediment, or suspended solids is not always linear with the concentration of chemical or soil and may vary considerably with the particular solid used. A number of authors (Carlberg and Martinsen, 1982; Matsuda and Sehnitzer, 1971; Ogner and Sehnitzer, 1970) have reported that soluble soil humic material associates strongly with phthalate esters, increasing their apparent water solubility and decreasing their apparent degree of soil sorption. Russell et al. (1985) reported that biodegradation in unsterilized soil-water systems affected the results of phthalate ester soil sorption studies.
A number of authors have published soil or sediment and water partition coefficients (Table 4). Banerjee et al. (1985) reported organic carbon-normalized partition coefficients (K~s) of DMP measured at different depths in a soil core ranging from 80 to 360. Russell and McDuffie (1986) determined soil partition coefficients for a number of phthalate esters. Measured K~s ranged from 69 for DEP to 87,000 for DEHP. Glcdhill et al. (1980) showed a BBP I ~ of 9,000 while Russell and McDuffie (1986) showed a comparable K~ for BBP of 17,000 (Table 4). Sullivan et al. (1982) showed a wide range ofDEHP K~s, from 12,000 to 1 million. More recently, Williams et al. (1995) presented sediment sorption coefficients for a number of sediments and various phthalate esters (DHP, DEHP, DIDP, and DTDP). Three sediments with organic carbons ranging from 0.15% to 1.88% organic carbon were investigated. So.rption results followed the Freundlich model. Average organic carbon normalized partition coefficients (Koc) ranged from 52,600 for DHP, 482,000 for DEI-IP, 286,000 for DIDP, and 1.2E+6 for DTDP. Williams et al. (1995) also noted an inverse dependence of experimental K~s with increasing solids concentration. The authors calculated particle-corrected I ~ values ranging from one to three orders of magnitude higher than the uncorrected experimental I ~ values. It was concluded that failure to consider particle effects with compounds such as phthalate esters, would result in significantly overestimating truly dissolved sediment porewater concentrations.
Several authors have examined the dissolved versus suspended particulate-bound fraction ofphthaiate esters in surface water samples. Germain and Langlois (1988) collected surface water samples from the St. Lawrence River. Both dissolved phthalate ester and phthalate ester bound to suspended particulate matter-bound (SPM) were analyzed. The SPM concantration was estimated to be 3.0 mg/L. About 14% of the total DnBP concentration was sorbed to SPM and 86% was dissolved. For DEHP, 53% was SPM-bound while only 47% was dissolved. No other phthalate ester had detectable SPM-bound fractions. Ritsema et al. (1989) separated SPM by centrifugation from surface water samples collected from Lake Yssel and the Rhine River (Netherlands). SPM ranged from 4-100 mg/L. Based on the geometric mean of th© range of SPM values, 98% of the DBP present was dissolved while only 2% was SPM-bound. For DEHP, 33% was estimated to be dissolved while 67% was estimated to be SPM-bound. Preston and AI-Omran (1986, 1989) collected a series of surface water samples from the River Mersey Estuary in the UK. SPM concentration reportedly was a relatively high 1524 mg/L. For DnBP, the authors reported 66% to 86% of the total was dissolved, while 14%
680 to 34% was particulate bound. Weekly samples have been collected for several years from the source and mouth of the Niagara River connecting two Great Lakes between the US and Canada (Data Interpretation Group, 1986-1993). SPM-bound and dissolved concentrations for several phthalate esters were measured in this monitoring program. The reports indicate that SPM averages 7.83 mg/L in the river samples. Dissolved DnBP averages 83.5% while SPM-bound DnBP was 16,5%. For DEHP, 73% was dissolved and 27% was particulate bound. Furtmarm (1993) reported similar findings. For the low molecular weight phthalate esters (DMP to Bttl'), 15-17% was particulate bound while for DEHP and Dnt)P 53-74% was particulate bound. The amounts of dissolved DEHP determined in these studies are consistently less than predicted from simple two-phase partitioning models. The differences are attributed to, in part, separation techniques that do not distinguish chemical complexed to organic colloids from truly dissolved phase chemical. A summary of partition coefficients between SPM and natural waters obtained from these studies are presented in Table 4.
It is generally accepted that only the truly dissolved phase of a non-polar organic chemical is bioavailable (Steen et al., 1980; McCarthy, 1983; Rodgers et al., 1985; Staples et al., 1985; McCarthy and Black, 1988; DiToro et al., 1991). Therefore, the partitioning of phthalate esters to colloidal and particulate organic carbon should be considered in analysis of field samples. However, most historical measurements are based on total concentrations that fail to differentiate free and complexed forms. Consequently, available field data may significantly overestimate bioavailability especially for the more hydrophobic phthalates that are expected to exist principally in the environment as complexed forms.
ENVIRONMENTAL TRANSFORMATIONS
Hvdrolvsis Phthalate esters are susceptible to hydrolysis, however at slow rates. The products of hydrolysis are an acid and an alcohol. Phthalate esters can undergo two hydrolytic steps, producing first the mono-ester and one free alcohol moiety and a second hydrolytic step creating phthalic acid and a second alcohol. Ester hydrolysis may be either acid or base catalyzed, with, in some instances, metal ions, anions, or organic materials serving as catalysts (Harris,
1982a). Phthalate esters are hydrolyzed at negligible rates at neutral pH (Table 5). Acid hydrolysis of phthalate esters is possible, but is estimated at four orders of magnitude slower than alkaline hydrolysis rate constants (Mabey et al., 1982). As an example of the limited extent of acid hydrolysis, Thuren and Sodergren (I 987) used extraction into
681
Table 5. Summary of Abiotic Degradation Half-lives of Phthalate Esters.
• b
Phthalate Ester
Aqueous Hydrolysis Half=Lives (years)'
Atmospheric Photooxidation Half=Lives (days) b
DMP
3.2
9.3-93
DEP
8.8
1.8-18
DAP
--
0.04-0.4
DPP
--
0.9-9.0
DnBP
22
0,6-6,0
DIBP
--
0.6-6.0
BBP
>0.3
0.5-5.0
DHP
--
O.4-4.0
DnOP
107
0.3-3.0
BOP
--
0.4-4.0
610F
--
0.2-4.0
DEHP
2000
0.2-2.0
DIOP
157
0.3-3.0
DINP
--
0.2-2.0
DIOP
--
0.2-4.0
711P
--
0.2-4.0
DUP
--
0.2-2.0
DTDP
--
0.2=2.0
Hydrolysis rate constant to monoester at pit--7, 25 C (Wolfe et aL 1980a), except for BBP (Gledhill et al., 1980) and DnOP (Howard, 1991). Predicted half-lives obtained by the Atmospheric Oxidation Program (Atldnson, 1988).
hydrated sulfuric acid to extract phthalate esters in a clean-up step for analysis and obtained 100% recoveries. Wolfe et al. (1980a) measured alkaline hydrolysis rate constants for a number of phthalate esters from free energy relationships. Hydrolysis rates decrease and corresponding half-lives increase with increasing alcohol chain length. Half-lives ranged from about 3 years for DMP to 2000 years for DEHP. Gledhill et al. (1980) estimated a hydrolysis half-life of>100 days for BBP. Hydrolysis is unlikely to be an important fate process for phthalate esters under typical environmental conditions (Schwartzenbach et al., 1992).
682 Photodegradation Aqueous photolysis occurs through absorption of UV light from sunlight in the region of 290-400 run. Shorter wavelengths are attenuated by passage through the atmosphere and water column. Longer wavelengths lack sufficient energy to break covalent bonds (Harris, 1982b). Photolysis can be mediated via either direct or indirect mechanisms. The mechanism of photolysis may be either through direct absorption of UV radiation by the chemical or by absorption of UV radiation by natural substances such as water with the formation of activated species such as singlet oxygen or hydroxy radicals that react with phthalate esters.
Few studies on phthalate ester photolysis are available. Gledhill et al. (1980) reported that a 1 mg/L solution of BBP exposed to sunlight for 28 days resulted in less than 5% degradation. It was concluded that BBP has an aqueous photolysis half-life of>100 days. Wolfe et al. (1980b) estimated a maximum near-surface phthaiate ester photolysis half-life of 144 days for phthaiate esters. This estimate was based on unpublished experiments with DMP spiked into surface water that was irradiated with sunlight for one week. Photooxidation ofphthalate esters in sunlit surface waters does not appear to represent an important transformation process. Based on the data of Wolfe et al. (1980b), Howard (1991) estimated that aqueous photooxidation half-lives range from 2.4 to 12 years for DEP and DnBP and from 0.12 to 1.5 years for DEHP.
In contrast to the minor role of photodegradation in natural waters, these reactions appear to be much more important in the atmospheric fate of phthaiate esters. Reaction with hydroxyl radicals is generally the most important photodegradation process for organic chemical pollutants in the atmosphere (Kelly et al., 1994; Atkinson, 1988; Meylan and Howard, 1993). Predicted photooxidation half-lives for phthalate esters based on hydroxyl radical attack were obtained from structure activity relationships contained in the Atmospheric Oxidation Program (AOP) (Meylan and Howard, 1993). Reported half-lives are specified as a range to indicate differences that are expected due to varying hydroxyl radical concentrations between pristine (3x10 "~radicals cm3) and polluted (3xI0 6 radicals cm-3) air. Results indicate that as alkyl chain length increases, susceptibility of phthalate esters to photooxidation also increases (Table 5). While limited experimental data are available for comparison with AOP predictions, one study reported an experimental atmospheric half-life for DEHP of about 1 day (Behnke et al., 1987), which agrees with model predictions.
a~air,~.gzl~a~ Biodegradation is a critical process affecting the environmental fate of phthalate esters. Considerable research has been conducted on the biodegradability of this chemical class over the last few decades. Microbes from diverse habitats have been shown to degrade phh'~alat¢esters and resulting intermediates (Stahl and Pessen, 1953; Mathur and Rouatt, 1975; Englehardt et al., 1975; Keyser ¢t al., 1976; Nagata et al., 1976; Klausmeicr and Osmon, [976; Kuran¢ ¢t al., 1977a, b; Englehardt et at., 1977; Englehardt et al., 1978; Nakazawa and Hayashi, 1978; Shimada and Shimahara, 1978; Yoshida
683 et al., 1979; Taylor et al., 1981; Benckiser and Ottow, 1982; Eaton and Ribbons,' 1982; Williams and Dale, 1983; Karegoudar and Pujar, 1984; Kurane, 1986; Gibbons and Alexander, 1989; Nomura et al., 1989; Yan et al., 1995). Representatives from both aerobic and some anaerobic environments include gram positive and gram negative bacteria and actinomycetes. Although some individual microbes are capable of completely mineralizing phthalate esters, more efficient metabolism appears to result from mixed microbial populations, typically found in the environment (Engelhardt et al., 1975; Aftring et al., 1981; Kurane, 1986).
Research suggests that the metabolic pathway for the microbial metabolism of phthalates under both aerobic and anaerobic conditions begins by ester hydrolysis to form monoester and corresponding alcohol. Under aerobic conditions, further enzymatic degradation of the monoester proceeds via phthalie acid by either a 3,5 or 4,5 dihydroxyphthalate pathway to procatechuate. Aromatic ring cleavage of procatechnate can then occur via either an ortho pathway that results in the formation of pyruvate and oxaloacetate or a meta pathway yielding a 13-ketoadipate that is further degraded to acetyl CoA and succinate (Eaton and Ribbons, 1982; Nomura et al., 1989; Kohli et al., 1989). Although less is known about the pathways of anaerobic catabolism, it appears that the monoester is degraded to phthalic acid and then further degraded by the same pathway used for benzoate (Afh-ing et al., 1981, Ejlertsson and Svensson, 1995). Benzoate has been shown to be readily degraded anaerobically (Battersby and Wilson, 1988; 1989; Shelton and Tiedje, 1984). Based on the recent comprehensive review by Ejlertsson and Svensson (1995), the proposed biodegradation pathway for phthalate esters is presented in Figure 1.
Numerous studies have been conducted to evaluate the primary and ultimate biodegradability of phthalates using both standardized and non-standardized test protocols. For the purpose of this review, studies can generally be divided into three types based on the nature of the test matrix: inoculum tests, freshwater and marine water tests, and soil/sediment tests. Inoculum tests involve spiking the test chemical to an aqueous solution containing inorganic nutrients and a microbial inoculum (e.g. wastewater effluent, sludge) obtained froma domestic or bench scale wastewater treatment system. The microbial inoculum used in such tests may or may not be acclimated to the test chemical. Acclimated inocula typically reduces or eliminates the time lag observed before the onset of biodegradation. Freshwater/marine and soil/sediment tests typically involve following the degradation of test chemical added to flasks containing natural water or soil/sediment. Alternatively, microcosm studies have also been used to assess the potential biodegradability of phthalate esters in natural waters and sediments. Microcosm studies have been performed with DEP (Lewis et al., 1985), DnBP (Tagatz et al., 1986); BBP (Adams et al., 1989; Adams and Saeger, 1993); DnOP (Sanborn et al., 1975) and DEHP (Sodergren, 1982; Perez et al., 1983; Davey et al., 1990). Field studies have also been conducted to assess phthalate ester biodegradation in soil (Falrbanks et al.~ 1985; Schmitzer et al., 1988). Such tests share the common
684
[~C 0 0 R COOR Dialkyl Phtha[ate
[~CO
0R COOH
Mono Alkyl Phthalate
Ic°? H
~:oo. or
I
OH
OH 3.4 Oihydroxy Phthalate
~'~/COOH A
=CO CoA I COOH
O
OH 4,5 Dihydroxy Phthalate
I
Phthalic Acid
ATP l + Co AADP
v'~COz OCoA
I
Benzoyl - CoA
otochatechuate
mete i ortho I cleavage
~1~0 0 H CHO /HOOC 2-hydroxy4-carboxym uconic sere leldehyde
oA r ' 1 Cycionex. 1. erie c a r b o x y l - CoA
I
~j~O0 H
U , ~ COOH COOH 13.Carboxy-CisCis-Muconate
[ . ~ (~1-~ 2-H y~roxycyclohexane LVJ Co,~arboxyl - CoA
3 Acetate + ~ 3Hz + CO2
¥
Acetate ÷ CO z + Succinate
T
.
.
.
.
CO CoA 11 CO CoA L~ Pimeiyl di.CoA
2 Pyruvate + CO z
Figure 1. General biodegradation pathway for phthalate esters in the environment.
685 objective of assessing the potential of indigenous microbial populations to degrade the test chemical. Consequently, an inoculum is not added nor is the influence of acclimation on the microbial consortia considered in these studies. Inorganic nutrients, however, are sometimes added.
An extensive summary of biodegradation test results for phthalate esters is provided in Table 6. Test conditions including the use of acclimated inocula, nutrient addition, temperature and initial phthalate concentration are provided for comparison. Tests assessing primary degradation are based on measuring the disappearance of parent phthalate by a specific analytical method (e.g. GC or HPLC analyses). In comparison, ultimate biodegradation is assessed by measuring carbon dioxide evolution or oxygen uptake under aerobic conditions or methane evolution under anaerobic conditions. For a number of studies compiled in Table 6, radiolabeled compounds were employed.
Aerobic degradation studies using inoculum tests have been performed by Saeger and Tucker (I 973 ); Barth and Bunch (1979); Urushigawa and Yonezawa (1979); Tabak et al. (1981); Patterson and Kodukula (1981); Sugattet al. (1984); O'Grady et al. (1985); Nyholm (1990); Struijs and Stolenkamp (1990); Aichinger et al. (1992); Association of Plasticizer Industry Japan (1994); Scholtz (1994a,b); and Exxon Biomedical Sciences, Inc. (1995a, b, 1996). Primary degradation for the lower molecular weight phthalates DMP, DEP, DBP and BBP occurred rapidly, typically exc~ding 90% degradation within a week even if unacclimated inocula were used. Most higher molecular weight phthalate esters demonstrated primary degradation in excess of 90% al~er 12 days using acclimated inocula. Slower rates of degradation we~ observed for unacclimated test systems. Ultimate biodegradation data for various phthalate esters are plotted in Figure 2. These results are based on 28-day experiments except for the data of Aichinger (1992). For this study, biodegradation was calculated fi'om the estimated, time-independent, growth yield obtained using a kinetic model. The studies by Saeger and Tucker (1976); Sugatt et al. (1984); Nyholm (1990); Struijs and Stolenkamp (1990); Scholtz (1994a, b); and Exxon Biomedical Sciences, Inc. (I 995b), measured CO2 evolution endpoints, whereas the studies by Aichinger et al. (1992); Association of Plasticizer Industry Japan (1994); and Exxon Biomedical Sciences, Inc. (1995a, 1996) measured oxygen uptake.
Figure 2 shows that aerobic biodcgradation test results for different phthalatc esters obtained from different studies are generally in good agreemcm. Consistent differences in results using tests with acclimated and unacclimated test inoculum are not obvious. However, results for DEHP reported by Struijs and Stolenkamp (1990) appear to be inconsistently lower than other studies. This discrepancy may be explained by the method in which DEHP was added to the test system. In this study, DEHP was impregnated into filter paper which was then introduced into the test system. As a result, the bioavailability of impregnated DEI-IP is likely limited by the desorption rate of DEHP from the filter paper. If the result from this study is excluded, all phthalate esters investigated undergo 260% ultimate
AS; WW SL
Aerobic-Ultimate
SG(4); SM SG SL SD (2) SG (2) $D(2)
Anaerobic-Primary
FW FW FW FW ,SL SL(2); SM SG(3); SL(3) FW WW FW WW SL(2); SM SL, WW
SM; AS SM; WW SM; WW SM; AS; SL
AS AS SM; AS
Aerobic-Primary
Test Matrix
N/N N/N
N/Y N/N N/N N/N N/Y N/N
Y/Y N/Y Y/Y Y/Y N/Y N/NR Y/Y N/N N/N N/N N/N N/N N/N N/Y N/N Y/Y N/N N/N N/N N/Y
Acclimation/ Nutrients
25 20
35 37 30 35 35 35
NR 25 22 22 25 NR 22 20 4 NR ? 30 25 NR 20 25 25 NR 25 20
Cone. (ppm)
100 27
30-81 0.5-10 500 81 20 50
0.1 - 1.0 20 I-3 3 5 5 20 0.0035 0.0035 0.0003 ? 500 1 703-1500 0.00074 0.00074 0.0018 NR 1 0.020
Dimethyl Phthalate
(*c)
Temp
Initial
28 3.75
7-70 7 30 56-96 70 56-96
1 7 1 1 7 7 28 3 I0 1 14 10 4 76-97 5 17 7 7 4 2
(Days)
Test Duration
Table 6: Summary of Phthalate Biodegradation Literature
90-98 50
100 77' 63 18-40 >70 18-40
100 48" >81 >90 100 >90 >99 100 0 97 ,,b 20-100 91 85 - 100 64-94 95 64 100 100 86-100 99.8
(%)
Degradation
15 16
17 18 10 13 34 13
t 2 3 3 4 5 6 7 7 8 9 10 11 12 14 14 14 14 11 71
Reference
O~ O0 O~
N/Y N/N Y/Y N/N N/N N/N N/Y
SG(3); SL(3)
FW WW FW WW SL(2); S M SL; WW
SC_~4);SM
N/Y
25 NR 22 20 4 20
N/Y N/NR Y/Y N/N N/N N/N N/N N/N
35
20 25 25 NR 25 20
NR
15-32
25 22
Diethyl Phthalate
37 37 35 35 35
22 25 20
N/Y Y/Y
Anaeroblc-Primary
Initial Cone. (ppm)
20 - 50
0.00086 0.00086 0.0024 NR I 0.020
! 000
20 I-3 3 5 5 20 0.0025 0.0025 0.19 25 0.02-2.2
20 20 - 200 81 81 50
20 55 0.020
Dimethyl Pbthalatc (cont.)
Temp (°C)
AS SM; AS SM; AS SM; WW SM; WW SM; AS; SL FW FW FW FW SL
Aerobic-Primary
SG; SM SG; SM SG(2); SM SG; SM SG(3)
N/Y N/Y N/NR N/Y N/Y
Y/Y Y/Y N/Y
SM; AS; SL SM;MI SL; WW
Anaerobic-Ultimate
Acclimation/ Nutrients
Test Matrix
7 - 70
5 17 7 7 41 2
287
7 1 ! 7 7 28 3 I0 ! 14 5 I- 129
70 140 7 77 32 - 56
28 NA 4
Test Duration (Days)
Table 6: Summary ofPhthalate Biodegradation Literature
100
95 0 99 98 86-100 100
95
36" >94.8 >90 ! 00 > 90 >99 100 0 76- l 0(P 14-100 19-99
82 94 - 100 58 - 88 41 58 - 88
86 96 c 99.6
(%)
Degradation
17
14 14 14 14 II 71
!2
2 3 3 4 5 6 7 7 24 9 25
34 21 22 23 17
6 19 71
Reference
O, oo
Aerobic-Primary AS SM ; AS SM; AS SM; AS; SL SM; WW SM; WW FW FW FW FW FW FW FW MW FW FW; SD FW
SG(3) SC
Anaerobic-Ultimate SG; SM SG(2); SM
SL; WW
Y/Y Y/Y Y/Y N/Y N/NR N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N
N/Y
N/Y N/NR N/Y N/Y
Y/Y Y/Y N/Y
N/N N/Y N/Y
SG SG(2) SC
Aerobic- Ultimate SM; AS; SL SM;MI
Acclimation/ Nutrients
Test Matrix
20-200 77 50 50-250
20 37 0.020
0.5 - 10 20 50-250
20 1-3 3 20 5 5 1 0.05 0.00325 0.00325 0.001 25
< 1.6 < 3.8 0.4-0.55
25 22 22 22 25 NR 25 NR 20 4 NR
28 28 25
Di-n-Butyl Phthalate
37 35 35 37
22 25 20
37 35 37
Diethyl Phthalate (cont.)
Temp (*C)
Initial Cone. (ppm)
7 ! 1 28 7 7 4 3 7 10 I 14 I0 10 7 7 9-11.5
17 28-56 32-70 100
28 NA 6
7 70 60- 100
Test Duration (Days)
Table 6: Smmnm~ o f Phthalate Biodegradation Literature
60 > 95 a 90 90 i00 a90 99 > 90 100 0 82" 33-100 45-95 55 49 32-49 80-100
33-56 0-32 0-76 5-90
95 93 ¢ 99.2
64' >90 85-100
(%)
Degradation
2 3 3 6 4 5 26 27 7 7 8 9 28 28 29 29 30
21 22 17 75
6 19 71
I8 34 75
Reference
O~ OO Oo
FW; SD MW M W ; SD MW SL(2); S M SD SD SL SL(2) SL(2) SL(2) SL SG(3); SL(3) FW WW SL; WW SL; GW
Aerobic-Ul~mme SM; AS; SL
SL SC SL; GW
SDO); SM
SG; SM SO (2) SO (3) SO; SM SO SG;SM
Y/Y
N/Y N/Y N/Y N/Y N/N N/Y N/Y N/N N/N N/N
N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N N/Y N/N N/N N/Y N/N
Matrix
Anaerobic-Primary
Acclimation/ Nutrients
Initial Conc. (ppm)
20 10
22
20
20 20 20 50 0.5-10 56 56 500 500 5
NR 0.024 5
NR
35 35 35 35 37 30 30 30 ! 8-68 10
0.4-0.55 0.3-0.5 0.3-0.5 0.3 I ! 0- 1000 10- 1000 500 200 800 1600 0.02-2.2 265-596 0.0016
25 25 25 25 25 28.5 28.5 30 21 2! 21 15-32 NR 25
Di-n-Butyl Phthalate (cont)
Temp (°C)
28
70 28 17-70 68 7 63 22-42 30 30 35-166
2 15
7
7-9 7-17 2-13 1 3 42-56 42-56 !0 182 182 182 34-129 76 7
Test Duration (Days)
Table 6: Summary of Phthalate Biodegradation Literature
57
100 100 80-100 100 83" 81 > 95-100 66 79 50~
99.3 50~
75
!00 95-100 I00 ! O0 97-100 59-96 52-89 92d 66-98 35-43 40 88-97 51-79 59
(%)
Degradation
6
34 34 17 i7 18 35 35 10 36 48
7I 48
14
33 33 25 12 14
33
30 30 30 3I II 32 32 I0
Reference
,4D
Aerobic-Primary SM; AS; SL SM; AS SM; WW SM; WW FW FW FW FW FW; SD FW; SD FW
SL; SM SL; WW
sG(12); SM
SG; SM SG; SM SO(2); SM SG(2); SM
A naerobic- Ultimate
FW; SD FW; SD FW; SD FW; SD SL; WW
SL(I); WW
N/N N/N N/N
Y/Y N/Y N/Y N/NR N/N N/N N/N
N/Y N/Y N/NR N/Y N/Y N/N N/N
N/Y Y/Y N/N N/N N/N N/N N/N N/N N/N N/N N/Y
SM; AS
SM;MI SL SL(4) SL(I)
Acclimation/ Nutrients
Test Matrix
20 20-100 72 72 50 1 1000
22 8 I 1000 1000 1000 0.082-8.2 NR NR NR 0.024
22 NR 25 NR NR NR 20 4 NR 20 20
28 I 7 7 9 7 7 10 7 2
5
0.00162
70 140 14-21 60 56 30 80
28 NA 30 80 80 53 14 14 14 14 12
Test Duration (Days)
20 NR 5 5 1 I 0.0035 0.0035 1 0.01-0.1
Butylbcnzyl Phthalate
35 37 35 35 35 22 23
21 25 22 23 30 4 22 12 28 5 20
Di-n-Butyl Phthalatc (cont.)
Temp (°C)
Initial Cone. (ppm)
Table 6: Summary of Phthalate Biodegradation Literature
100
87 96 100 a 90 99 100 100 0 95 47-60*
80 100 32-85 24-59 3 2 - 85 98 84
81 92 c 97 84-99 83 2 71-85 80 95 50 99.8
(%)
Degradation
14
6 42 4 5 42 43 7 7 44 45
34 21 22 41 17 38 39
37 19 38 39 39 39 40 40 40 40 71
Reference
'*,D 0
SM; AS
Aerobic-Primary
SG(2); S M FW; SD
so(2); S M
SO; SM SM;SG?
Anaerobic-Ultimate
SM; AS; SL SM; AS SM;MI FW; SD FW; SD FW; SD SM; AS AS; WW
Aerobic-Ultimate
SG; SM SG; SM SG; SM SG SG SG(3) SD(3); SM FW; SD
Y/Y
N/Y NR/NR N/Y N/Y N/N
Y/N Y/Y Y/Y N/N N/N N/N N/N N/N
N/Y N/Y N/Y N/N N/N N/Y N/Y N/N
Y/Y N/N N/N
WW FW WW
Anaerobic-Primary
Acclimation/ Nutrients
Test Matrix
Initial Cone. (ppm)
22
Dihexyl Phthalate
35 NR 35 35 NR
22 NR 25 NR NR 20 N/R 25
35 30 35 37 35 35 30 NR
25 25 NR
1-3
20 NR 68 50 I
20 20 9-45 l 1 0.Ol-O.I N/R 100
20 62 50 4 20 20 62 I
0.00162 0.0067 NR
Butylbenzyl Phthalate (cont.)
Temp (°C)
1
70 28 28-56 32-56 28
28 27 NA 28 28 30 28 14
48 63 48 7 28 42-70 42- 100 2
i7 7 7
Test Duration (Days)
Table 6: Summary ofPhthalate Biodegradation Literature
> 93
I00 <10 0-24 0-100 45
43 95 91-93 ~ 50 51-65 10.4 96 81
100 52 100 66' 90 75-98 100 50'
95 100 I00
(%)
Degmdmion
3
34 44 22 17 46
6 42 19 46 44 45 44 15
34 35 17 18 34 17 35 46
14 14 14
Reference
O~
SG; SM
Anaerobic-Ultimate
SL; WW
SM;MI
Aerobic-Ultimate
SG; S M SG SO(3) SG
Anaerobic-Primary
FW FW FW SL SC FW WW
SM; WW SM; WW
AS
Aerobic-Primary
SM; AS; SL SM; AS
N/NR
Y/Y N/Y
N/Y N/N N/N N/N
N/Y N/Y N/NR YINR N/N N/N N/Y N/N N/N N/N N/N
Y/Y N/Y
Y/Y N/Y N/N
SM; AS (cont.) SM; AS; SL AS FW
Aerobic-Ultimate
Acclimation/ Nutrients
Test Matrix
35
25 20
35 37 35 35
25 25 NR NR 20 4 26 15-32 18-69 25 NR
20 35
3 20 20 25
68
8-39 0.020
50 4 20 20
20 5 5 5 0.0025 0.0025 2.5 0.02-2.2 500 0.0034 NR
Di-n-Oetyl Phthalate
22 22
22 25 ?
DihexylPhthalate(cont.)
Temp (°C)
Initial Cone. (ppm)
56
NA 20
70 32 70 70
7 7 7 7 10 I0 5 34-96 30 7 7
28 28
I 28 7 14
Test Duration (Days)
Table 6: Summary of Phthalate Biodegradation Literature
0
85-92 c 52.3
5 0 12-44 30
10~ 0 0 92 85 0 50 "d 80 31 100 97
77 80
a 90 > 99 38" 30-95
(%)
Degradation
22
19 71
17 18 17 34
4 5 5 7 7 47 25 36 14 14
6 70
3 6 2 9
Reference
tO
SM; AS SM; AS SM; AS SM; AS SM; AS SM; AS; SL SM; WW SM; WW SM; WW SM; WW FW
Aerobic-Primary
SM; AS; SL
Aerobic-Ultimate
SM; AS SM; AS SM; AS; SL
Aerobic-Primary
SM; AS; SL
Aerobic-Ultimate
SM; AS SM; AS SM; AS; SL
Y/Y Y/Y Y/Y Y/Y Y/Y Y/Y N/Y Y/Y N/NR Y/NR N/N
Y/Y
Y/Y Y/Y Y/Y
Y/Y
Y/Y Y/Y Y/Y
Initial Conc. (ppm)
50 20
22
22 22 22
20
1-3 3 20
Butyl2-Ethylhexyl Phthalate
35 35
Di-n-OetylPhthalate(cont.)
Temp (°C)
I-3 3 20
22 15 NR 22 22 22 25 25 NR NR NR
3.3-33 3.3-33 3.3 I-3 3 20 5 5 5 5 l
22 20 Di (2-Ethylhexyl) Phthalate
22 22 22
Di (n-Hexyl, n-Octyl, n-Decyl) Phthalate
N/Y N/Y
SG(14); SM SG; SM
Aerobic-Primary
Acclimation/ Nutrients
Test Matrix
I 1 1 I 3.5 28 7 7 7 7 5
28
! 3 28
28
1 3 28
56-70 70
Test Duration (Days)
Table 6: Summary of Phthalate Biodegradation Literature
31-88 22-36 74 81.5 ~ 90 > 99 0 75 0 95 40
87
60 ~ 90 > 99
90
60 ~ 90 > 99
0-13 13
(%)
Degradation
49 49 43 3 3 6 4 4 5 5 42
3 3 6
6
3 3 6
17 34
Reference
~O
SG; SM SG SG(2) SG $L SD(2); SM SD LL SL (2)
N/Y N/N N/N N/N N/N N/Y N/N N/Y N/N
N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N N/Y N/Y N/N N/Y N/Y N/Y N/Y N/Y N/N N/N N/N N/N N/N
FW FW FW FW FW FW MW MW MW SL; SM SL; SM SG(3); SL(3) SL SL; SC SL; SC SL; SC SL; SC SL; SC FW FW WW SL $L
Anaerobic-Primary
Acclimation/ Nutrients
Test Matrix
(Days)
(ppm)
30 37 35 35 30 30 NR 30 20
7.8-78 4 20 20 500 7.8-78 NR 78 1
25 1 0.05 5 20 0.0035 4 0.0035 --25 . . . . . . . . . . . . 18 0.001-0.1 1 0.001-0.1 NR 22,000 NR 22,000 NR 372-578 30 500 22 1230 22 4260 22 17700 22 21900 N/R 2000 N/R 0.05 25 0.0051 NR NR NR 5 NR 250
63 32 70 70 30 365 14 61 1O0
20 5 10 10 14 10 10 2 28 30 24-30 76-97 30 70 70 70 70 190 2 7 7 80 80
Test Duration
Initial Conc.
Di (2-Ethylhexyl) Phthalate (cont.)
Temp (°C)
Table 6: Summary o f Phthalate Biodegradation Literature
0 0 0-28 10 33 0-14 0 1 64-97
> 99 50 37 0 31-88 35-75 26 50 50 0 > 90 60-94 33-92 80-86 94 32-69 14-37 85 10 79 33 85 50
(%)
Degradation
35 I8 17 34 10 35 52 52 72
26 38 7 7 9 28 28 50 50 51 51 12 10 20 20 20 20 20 27 14 14 73 73
Reference
SL (2) SL (I) SL (I)
FW FW FW; SD SL; WW SL; SM SL;SM SL SL; SG SL; MI SL SL SL SL
SL(3)
AS; MW MW FW FW FW FW SL(3)
Y/Y Y/Y N/Y Y/Y N/Y N/Y N/Y N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N N/Y N/N N/Y N/N N/N Y/Y N/N N/N N/N N/N
N/N N/N N/N
Aerobic-Ultimate SM; MI SM; AS; SL SM; AS SM; AS SM; AS SM; AS SM; AS
Acclimation/ Nutrients
Test Matrix
Initial Cone. (ppm)
25 22 2! 20 20 22 25 25 <10 >20 29 29 29 29 N/R N/R NR 29 l0 20.5 NR NR 22 21-24 NR 22 NR NR 20-24
13.5 14.1 12.8
8-40 20 20 20 50 46 100 i00 0.002 0.002 I 0- 100 ppt 0.1 - I ppb 1- 10 ppb 10-100 ppb 0.5 10 0.1 20 ppt-200 ppb 1.4 0.1 NR NR I 2-20 l 0.029 1 2 2
! 1 1
Di (2-Ethylbexyl) Pbthalate (cont.)
Temp (°C)
28 28 27 28 28 28 14 14 14 40 40 40 40 28-63 63 32-34 28 27 28 14 14 30 146 20 33 111 446 7
NA
100 100 100
Test Duration (Days)
Table 6: Summary ofPhthalate Biodegradation Literature
38 61 12 55 59 86 50 22 93 98 2
0-80
4.5 • 63 80 29 11-16 33-69 71.2 42.3 34.9 62.3 43-58 22-34 4-28
82 86
90-95 c 86
33-89 96-99 29-72
(%)
Degradation
19 6 53 42 54 55 26 15 56 56 57 57 57 57 74 74 58 59 60 12 51 51 38 61 62 63 63 63 64
72 72 72
Reference
t~
SM; AS; SL SM; AS
Aerobic-Ultimate
SM;AS SM;AS SM; AS; SL
Aerobic-Primary
SG;SM SG;SM SG;SM SG;SM SG(3);SM SD
SG;SM
SG; SM SG;SM SG;SM SG;SM SG(2); SM
A naerobic- Ultimate
SC
SL (2)
FW; SD FW; SD SD SL (2)
Y/Y N/Y
Y/Y Y/Y Y/Y
N/Y N/Y N/Y N/Y N/? N/Y N/? N/? N/? N/? N/Y N/N
N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N NN
SL SL SL FW; SD
FW; SD
Acclimation/ Nutrients
Test Matrix
Initial Cone. (ppm)
22 25
22 22 22
20 100
1-3 3 20
20 20 100 200 68 68 34 68 136 271 50 I
2.6-14 0.1 0.1 0.018 0.018 0.018-10 0.018 NR ! 1 500
Di-lsooctyl Phthalate
35 37 37 37 35 35 35 35 35 35 35 22
15-30 22 35 5 12 22 28 20 20 13.5 18-69
Di (2-Ethylhexyl) Phthalate (cont.)
(*C)
Temp
28 28
I 4 28
70 140 140 140 56 77 60 60 60 60 32-70 3O
!i5 33 56 28 28 28 28 28 100 100 30
Test Duration (Days)
Table 6: Summary of Phthalate Biodegradation Literature
57 56
84.5 a 90 > 99
9 50 7 4 0 0 48 26 19 6 0-9 0
68 23 12 I 3 6-24 10 20 2.6-47 12-33 41
(%)
Degradation
6 68
6
13 13
34 21 21 21 22 23 41 41 41 41 17 38
65 66 66 40 40 40 40 67 72 72 36
Reference
~D
SM; AS; SL SM; AS Ski; AS
Aerobic-Primary
SM; SL; AS SM; AS
Aerobic-Ultimate
SM; AS; SL SM; AS FW FW WW
Aerobic-Primary
SM; AS; SL SM; AS SM; AS SM; AS FW; SD
Aerobic-Ultimate
SM; AS SM;AS SM;AS FW
Y/Y N/Y Y/Y
y/Y N/Y
Y/Y N/Y N/N N/N N/N
Y/Y N/Y N/Y N/Y N/N
Y/Y Y/Y Y/Y N/N
N/N
FW; SD
Aerobic-Primary
Acclimation/ Nutrients
Teat Matrix
Initial Cone. (ppm)
20 45 35 100 0.02-10 NR
20 3.3-13.3 1 0.00052 NR
22 NR 22
20 3.3-13.3 !-3
22 20 22 46 Di(Heptyl, Nonyl, Undecyl)Phthalate
22 NR NR 25 NR
Di-Undecyl Phthalate
22 22 22 25 22 28
22 22 22 25
I-3 3 20 1
NR 0.002-10 NR
Di-lsononyl Phthalate
12 22 28
Di-lsooetyl Phthalate(eonL)
Temp (*C)
28 1 1
28 28
28 I 5 7 7
28 28 28 28 28 28
I 4 28 12
28 28 28
Test Duration (Days)
Table 6: Summary ofPhthalate Biodegradation Literature
> 99 51 65
76 57
> 99 29-45 20 100 76
62 70 57 71 < 1-8 2
68 ~ 90 > 99 > 95
2 < 1-10 4
(%)
Degradation
6 43 3
6 55
6 42 42 14 !4
6 55 69 26 40 40
3 3 6 26
40 40 40
Reference
"..,I
• b ° d " (#) AS MI
Y/Y
Y/Y Y/Y
Y/Y N/Y N/Y N/N
Y/Y Y/Y
Y/Y Y/Y
Y/Y N/N
Temp (*C)
3 1
22
22 22
22 22 25 25 Di-Tridecyl Phthalate
22 22
20
20 I-3
20 48 100 100
20 1-3 3
22 20 NR 20 Di-lsodeeyl Phthalate
22 NR
Di (Heptyl, Nonyl, Undecyl) Phthnlate (cont.)
Acclimation/ Nutrients
Calculated from reported first-order half lives. Primary removal process assumed to be biodegradation. Calculated from reported kinetic parameters. Includes losses due to other potential removal mechanisms (e.g., sorption). Chemical impregnated into filter paper. # of test matrices investigated e.g., SL(4) indicates 4 soils were tested. Activated Sludge Microbial innoculum from batch reactor fed with test compound as sole carbon source.
SM; AS; SL
Aerobic-Ultimate
SM; AS; SL SM; AS
Aerobic-Primary
SM; AS; SL SM; AS SM; AS AS; WNV
Aerobic-Ultimate
SM; AS; SL SM; AS
Aerobic-Primary
SM; AS SM; AS
Aerobic-Ultimate
SM; AS (cont.) FW
Test Matrix
Initial Conc. (ppm)
28
28 1
28 28 28 21
28 1 9
28 27
3.5 5
Test Duration (Days)
Table 6: Summary ofPhthalate Biodegredation Literature
37
> 50 51
56 67 67 42
> 99 68 90
98 86
55
90
(%)
Degradation
6 55 26 15
6 42
3 42
Reference
OO
Fresh water (e.g., river, lake, pond) Landfill leachate Municipal Solid Waste Marine Water No / Yes Not Reported Synthetic Compost Sediment Sludge Soil Synthetic aqueous medium Wastewater
1. Barth and Bunch, 1979; 2. Urushigawa and Yonezawa, 1979; 3. O'Grady et al. 1985; 4. Tabak et al. 1981; 5. Patterson and Kodukula, 1981; 6. Suggat et al. ! 984; 7. Ritsema et al. 1989; 8. Zoetman et al. 1980; 9. Hattori et al. 1975; 10. Shanker et al. 1985; I 1. Russell et al. 1985; 12. Kinconnin and Lin, 1985; 13. Madsen et al. 1995; 14. Furtmann, 1993; 15. CITI, 1992; ! 6. Osipoff, 198 I. 17. Shelton and Tiedje, 1981 ; 18. Ziogou et al. 1989; 19. Aichinger et al. 1992; 20. Maag and Lokke, 1990; 21. O'Connor et al. 1989; 22. Horowitz et al. 1982; 23. Battersby and Wilson, ! 989; 24. Lewis et al. 1984; 25. Overeash et al. 1986; 26. Association of the Plasticizer Industry,Japan, 1994; 27. Sehouten et al. 1979; 28. Kodama and Takai, 1974; 29, Steen et ai. 1980; 30. Walker et ai. 1984; 31. Cripe et al. 1987; 32. Tagatz et al. 1986; 33. Oven,ash et al. 1982; 34. Shelton et al. 1984a; 35. Painter and Jones, 1990; 36. Snell Environmental Group, 1982; 37. Scholz, 1994a; 38. Johnson and Lulves, 1975; 39. Inman et al. 1984; 40. Johnson and Heitkamp, 1984; 41. Battersby and Wilson, 1988; 42. Saeger and Tucker, 1976; 43. Saeger and Tucker, 1973; 44. Gledhill et al. 1980; 45. Adams et al. i 989; 46. Adam and Saeger, 1993; 47. Sanborn et al. 1975; 48. Chaure~ et al., 1996; 49. Wooek, 1979; 50. Perez et al. 1985; 51. lrvine et al. 1993; 52. Schwartz et al., 1979; 53. Seholz, 1994b; 54. Struijs and Stolenkamp, 1990; 55. Exxon Biomedical Sciences, Inc., 1995a; 56. Davey et al. 1990; 57. Subba-Ran et al. 1982; SS. Wylie et al. 1982; 59. Rubin et al. 1982; 60. Sodergren, 1982; 61. Fairbanks et al. 1985; 62. Efroymson and Alexander, 1994; 63. Schmitzer et al. 1988; 64. Kloskowski et al. 1981; 65. Aranda et al. 1989; 66. Scheunert et al. 1987; 67. Fish et al. 1977; 68. Nyholm, 1990; 69. Exxon Biomedical Sciences, Inc., 1995b; 70. Exxon Biomedieal Sciences, Inc., 1996b; 71. Ye Changming and Tian Kang, 1990; 72. Rudel et al., 1993; 73. Kirehman et al., 1991; 74. Dorfler et al., 1996; 75. Ejlertsson et al., 1996.
References;
FW LL MSW MW N/Y NR SC SD SG SL SM W%V
Table 6: Summary of Phthalate Biodegradation Literature (cont.)
o~
700
I
I
I
I
I
DTE~ DUP
I
I
I
I
rl
0
O • ¢~ 711P
•
•
QD
DEeP
-J
0
COQ n
(3
~. el0P 0
BOP
<
mr
O r, "=l)-
BlIP
0 m• I
0
I
10 20
I
I
I
I
I
I
80
40
50
60
70
80
I-
I
O0 100
Ultimate Blodegradatlon
Figure 2. Ultimate biodegradation of phthalate esters in standardized tests based on CO2 evolution (circles) or 02 uptake (squares). Filled symbols indicate acclimated inoculum was used.
degradation within 28 days, with the exception of DTDP (Figure 2). The extent of ultimate degradation also tends to decrease as the size of the alkyl chain increases. Anaerobic inoculum tests have been conducted by Gledhill et al. (1980); Shelton and Tiedje (1981); Horowitz et al. (1982); Shelton et al. (1984); Shelton and Tiedj¢ (1984); Battersby and Wilson (1988, 1989); O'Connor et al. (1989); Ziogou et al. (1989); Painter and Jones (1990); and Madsen et al. (1995). Results of primary and ultimate biodegradation ~tudies under anaerobic conditions tend to be more variable than aerobic tests (Table 6). However, several generalizations can be made. First, the rate of degradation under anaerobic conditions is retarded relative to aerobic tests. Second, a similar, but less well-defined trend is observed between alkyl chain length and the extent of anaerobic degradation. Increased variability in anaerobic biodegradation
701 suggests that the nature of the inoculum influences test results. This is supported by the work of Horowitz et al. (1982) who found that the source of sludge had a pronounced effect on the rate of anaerobic biodegradation of selected phthalate esters.
Primary biodegradation half-lives of phthalate esters in natural waters were estimated from data in Table 6 assuming first order kinetics. For studies in which test duration or percent biodegradation were reported as ranges, average values were used to compute mean half-lives. For studies in which 100% degradation is reported, half-lives were estimated by dividing the test duration by three. Although such an approach is simplistic, this analysis provides a convenient method to normalize the extensive biodegradation database assembled. Data used to construct Figure 3 are based on observed biodegradation in unacclimated test systems, using fresh or marine waters, with or without sediment, under aerobic (or partially aerobic) conditions. Studies by Perez et al. (1985), Ritsema et al. (1989), and Furtmann (1993) suggest primary biodegradation rates are significantly retarded at low temperatures. Therefore, tests performed at low temperatures (e.g. < 5"C) were excluded in this analysis. Half-lives shown in Figure 3 range from <1 day to about 2 weeks. Clear trends between different phthalate esters or between freshwater and marine waters are not apparent. Limited data for DBP and BBP suggest that addition of sediment does not significantly influence primary biodegradation rates.
Research on the ultimate biodegradability of phthalate esters in natural waters has focused principally on DEI-IP. Using biodegradation test results reported by Wylie et al. (1982) with unfiltered water samples, a mean DEHP half-life of 78 days is calculated. Other studies by Subba-Rao et al. (1982) and Rubin et al. (1982) yield half-lives for DEI-IP ultimate biodegradation in eutrophic lake water ranging from 12 to 64 days. In contrast, no DEHP mineralization was observed after 60 days in water from an oligotrophie lake (Rubin et al. 1982). The lack of mineralization in the nutrient Ix)or lake water was attributed to either the absence of competent microbial populations or the lack of growth factors required by indigenous degraders. In a marine microcosm study by Davey et al. (1990), the ultimate degradation half-life for DEHP averaged 12 and 67 days, respectively at temperatures <10 and >20 C, respectively. Comparison ofth6~¢ results to primary biodegradation half-lives for DEHP (Figure 3) suggests ultimate biodegradation half-lives are about art order of magnitude higher.
Half-lives in soils for the more well studied phthalate esters were calculated using data in Table 6 and the same approach described previously (Figure 4). Data used for this analysis were based on unacclimated test systems that did not exclude oxygen, but in many cases, likely reflect only partially aerobic conditions. Results repgrted by Kloskowski et al. (1981) for DEHP were not used due to the short time frame of this experiment (7 days). Figure 4 includes soils half-lives which exhibit considerable variation ranging from about 1 to 75 days for primary and 4 to 250 days for ultimate biodegradation. Such variation likely reflects differences in the biodegradation test protocols employed and
702
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o
o
o
ao
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ago
o
o
o
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,
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1.0
,
,
,
,,
r-
10.0
100.0
Prlm~-y Degra,'~tlon I-mr-Life (,~,,ye)
Figure 3. Primary biodegradation half-lives of phtbolate esters in fresh (circles) and marine (squares) waters under aerobic conditions. Filled symbols denote results for test systems containing sediment.
D~'IP
~D
DSP
•
DEP
eO
Q
•
,
,
,
....
0
e6
•
DMP
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,
10.0
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,
1000
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1000.0
De~'amam H~a~Je emy~
Figure 4. Primary (open symbols) and ultimate (filled symbols) biodegradation half-lives of phthainte esters in sofia.
703 in the microbial populations associated with the various soils investigated.
Limited data are available on
biodegradation in sediments. Based on lab and field studies, Tagatz e( al. (1986) reported primary biodegrad~tlon rates in sedimems of 3=4 weeks for DBP. In an earlier lab study, Fish et al. (1977) provide primary biodegradation data indicating a half-life for sediment-bound DEHF of ca. 3 months. In an anaerobic test system, Madsen et al. 0995) found that ultimate biodegradation of DMP and DIBP added to either a freshwater swamp or marine sediment was <30% after 56 days. Johnson & Heitkamp (1984) showed that the ultimate biodegradatinn of DIOP and DINP in sediments was generally <10% after 28 days. These studies suggest that primary biodegradation half=lives are on the order of several months with ultimate biodegradation proceeding more slowly with half=lives that may exceed I00 days. Since oxygen may be limiting in soils/sediments, the longer half=lives observed in ihese matrices when contrasted to natural waters (Figure 3) are consistent with the limited extent of biodegradation observed in anaerobic inoculum tests discussed previously.
Several generalizations are apparent from our review of phthalate ester biodegradation literature. First, phthalam esters may be used by aerobic and anaerobic microbes as a source of carbon and energy. Second, phthalate esters are degraded in both standardized biodegradation tests and in tests simulating different environmental compartments. The extent of biodegradation occurring over the course of these tests suggests that phthalate esters are not expected to be highly persistent in most environments as is sometimes incorrectly asserted. Third, primary degradation in water, sediment and soil coropartments is expected to be controlled by biodegradation rather than abiotic loss mechanisms. Fourth, based on available data for unacclimated test systems, the primary degradation half-life in water is expected to be on the order of less than one week, while the half-lives in soils ranges from less than one week to severalmonths. These values are consistent with estimates obtained by several recent independent reviews on phtbalat~ ester bio~gradafion (Peijnenburg et al., 1991, Howard 1991; Boethling et al., 1995). Lastly, longer half-lives are likely under anaerobic conditions and in cold, nutrient poor environments.
BIOACCUMULATION
Bioaccumulation refers to the accumulation of contaminants in the tissue of biota via all exposure routes (i.e., water + diet) whereas bioconcenu'ation refers to accumulation due to aqueous exposure alone (Veith et al., 1979a). These processes are quantified by the bioaccumulation and bioconcentration factor, respectively (i.e., BAF or BCF). The BCF and BAF define the ratio of concentrations in tissue to that in water and am expressed in units of mL per gram body weight, on a dry, wet or lipid basis. The BAF for a chemical can be derived based on field data using measmed water and tissue levels or can be predicted using bioaccumulation models if sufficient toxicokinetic ~
are available
(Tbomann, 1989; Clark et al., 1990; Oobas, 1993). In contrast, the BCF is derived from a laboratory test and ¢xtaesses the partition coefficient of a chemical between test organism and water at steady-state. In most oases, bioc,omentmfion
704 can be adequately described by simple first-order kinetic models so that the BCF may be calculated by dividing the uptake clearance, K , in units ofmL g-t day-~,by the elimination rate, K2, in units of day ~.
Numerous laboratory studies have investigated the bioconcentration, and to a lesser extent, bioaccumulation of phthalate esters. Many investigators have used radiolabeled phthalate esters due to the greater sensitivity and convenience afforded by radioisotopes. For chemicals such as phthalate esters that are subject to metabolism, BCFs based on total radioactivity are expected to be higher than BCFs based on parent compound since radiolabeled metabolites can contribute significantly to total radioactivity. A comprehensive compilation of phthalate ester bioconcentration studies with fish and other aquatic organisms is provided in Tables 7 and 8, respectively.
In cases where toxicokinetic parameters were not reported but raw data were provided, the original data were reanalyzed using the non-linear regression procedure included in SYSTAT (Wilkinson, 1990) to estimate these parameters and the resulting steady-state BCF. Results provided in Tables 7 and 8 include several studies in which organisms were fed or had the opportunity to feed (e.g., microcosm studies) on contaminated prey. Therefore, some of the reported BCFs may be indicative of BAFs due to dietary exposure.
In the microcosm studies by Metcalf et al. (1973), Sanborn et al. (1975) and Perez et al. (1985), BCFs were estimated by dividing the total ~4Cradioactivity in the various test organisms at the end of the experiment (30-33 days) by the average )4C activity in the aqueous phase during the test. For the first two studies, this average was based on measurements at the start and end of tests. Since total activity in the water column decreased by less than a factor of four in both studies this average provides a reasonably constant exposure concentration for calculating BCFs. In contrast, total activity dropped three orders of magnitude in a 27-day DEHP microcosm experiment reported by Sondergren (1982). Due to the large variation in exposure concentration it was judged that reliable BCFs could not be determined from this study.
Severul early investigations that used chromatographic techniques to distinguish parent phthalate esters from )(C metabolites, raised environmental concerns about phthalate esters as a result of the high BCFs reported (Metcalf et al., 1973; Sanborn et al., 1975; Sodergren, 1982). However, these values are misleading because calculations are based on parent phthalate ester concentrations determined in tissue and water samples at test termination only. To provide more realistic BCFs for parent phthalate esters that are derived from radiotracer studies, the BCF, based on total n4C, was multiplied by the percentage of the total radioactivity that was reported in tissue as parent compound at test termination.
Studies with fish (Table 7) indicate that Kt and BCF increase with increased t~-mporatureand decreased body weight (Karara and Hayton, 1989; Tart et al., 1990). For the more hydrophobic phthalate esters, K. and BCF tend to decrease
NR (-2.5)
0.09 ± 0.04
Pimephalespromelas
(Fathead Minnow)
(Sheepshead Minnow)
Cyprinodon variegatus
(English Sole)
Parophrys vetulus
(Bluegill Sunfish)
28.3 4- 1.0
blR (-l.0)
Lepomis macrochirus
(Bluegill Sunfish)
NR (~1.0)
0.37 - 0.94
Lepomis macrochirus
(Bluegill Sunfish)
Lepomis macrochirus
(End,fish Sole)
Parophrys vetulus
(Bluegill Sunfish)
32.7 ± 1.6
0.37 - 0.94
Lepomismacrochirus
(Sheepshead Minnow)
NR (-2.5)
0.37- 0.94
Weight (g wet)
Cyrinodon variegatus
(Bluegill Sunfish)
Lepomis macrochirus
Organism
25.4
N-R
12
?
NR
16
i2
16
NR
16
Temp (°C)
FT
S
FT
FT?
FT?
FT
FT
FT
S
FT
Expos.
ad. lib.
Yes
No
No
?
?
30 -200
9.4
Diethyl P h t h a l a t e
I00
8.7
5 50
100
[62l
40 - 81
6
19 - 39
.
.
165 145
11.7
[400]
.
.
225 - 457
Di-n-Butyl P h t h a l a t e
20 - 250
34
2
9.7
K, ,
.
.
(mL/g wet'ld I)
Butylbenzyl P h t h a l a t e ad. lib.
No
ad. lib.
No
Conc. ~g/L)
Exposure
Dimethyl Phthalate
Feeding
Test Conditions
.
.
0.08 0.07
. . . .
. . . .
0.34 - 0.69
. . . .
0.34 - 0.69
2,068 2,125
4494 453
293
1882
663
117
57
167 s 172 s
__
I
12
Total Parent BCF I BCF' (mL/g wet "l) (mL/g wet")
. . . .
0.34 - 0.69
(d "l)
K2 I
Table 7. Summary o f Bioconcentration Studies for Phthalate Esters with Fish.
Ref.
-,O C2~
! .24 ±
0.31 1.24 ±
0.31 1.24 ±
0.31 1.24 ±
0.31 ! .24 ±
0.31 1.24 ±
0.31
Pimcphalespromelas
(Fathead Minnow)
Pimephalespromelas
(Fathead Minnow)
Pimephalespromelas
(Fathead Minnow)
Pimephalespromelas
(Fathead Minnow)
Pimephalespromelas
(Fathead Minnow)
Pimephalespromelas
(Fathead Minnow)
(Blue[gill Sunfish)
Lepomis macrochirus
1.0 - 2.0
0.31
(Fathead Minnow)
22?
25
25
25
25
25
25
FT
FT
FT
FT
FT
FT
FT
FT
! .24 +
Pimephalespromelas
Gambusia sp. 25
S
26
NR (- 0.1)
(Mosquito Fish)
(carp)
FT
Expos.
?
Temp (*C)
Exposure Cone. 0zg/L) K, J (mL/g wet"d')
25
5% bdw
Yes
Yes
Yes
Yes
Yes
Yes
Yes
N Y
5.7 5.6
62
30
14
8.1
4.6
2.5
t.9
100:1: 19
--
767
8.2 [21
17 [6]
76 [9]
132 [161
80 [34]
82 [69]
707 [29] 9
2046
Di(2-Ethylhexyl) Phthalate
Yes
D i - n - B u t y l P h t h a l a t e (cont.)
Feeding
Test Conditions
?
Weight (g wet)
Cyprinus carpio
Organism
0.67 ~ 0.47
0.05, 0.04 I0.041
0.06, 0.02 [0.05]
0.25, 0.03 [0.05]
0.34, 0.05 [0.06]
137 + 6 I°
114+87
155 + 36
295 ± 65
304 + 51
384 + 25
817±72
900 + 69 0.10, 0.06 [0.06]
745 + 627
[0.05]' 0.09, 0.06 [0.11 ]
11
51
124
182
270
564
630
588
2
3.6
Total Parent BCF L BCF I (mL/g wet") (mL/g wet")
0.097, 0.048
--
K2' (d")
Table 7. Summary of Bioeoncentration Studies for Phthalate Esters with Fish.
lO
9
18
Ref.
Oncorhynehus mykiss,
(Carp)
Cyprinus carpio
(Mullet)
Mugil cephalus
formerly Salmo gairdneri (Rainbow Trout)
Oncorhynchus mykiss,
formerly Saimo gairdneri (Rainbow Trout)
30
4.3 :t: 0.4
2.9 61.3 441
25
24
12 12 12
I0 lO 10
NR (--O.1) NR
(-0.O
l0 16 23 29 35
NR
20-25
12
16
Temp (°C)
-5 -5 -5 - 5 - 5
1 1 1 1 !
Cyprinodon variegatus
(Sheepshead Minnow)
(Sheepshead Minnow)
NR (-2.5)
I -5
26.7 ± 1.3
0.37 - 0.94
Weight (g wet)
Cyprinodon variegatus
(Golden orfe)
Leuciscus idus melanotus
(English Sole)
Parophrys vetulus
(Bluegill Sunfish)
Lepomis macrochirus
Organism
FT
FT
S S S
FT FT FT
S S S S S
S
S
FT
FT
Expos.
Conc. (#g/L)
Exposure
No No
2% bdw
2 - 4% bdw
No No No
0.1-1.0
50
30 30 30
5 14 54
60
60 60
No
5% bdw 5%bdw 5% bdw
60 60
! 00 500
50
25 - 300
5.8
No No
No
No
No
ad. lib.
. . .
85 - 104
0.39
[2.3]
[3.71
[1.25]
[l.S]
. . .
[0.7] [0. l 1]
[0.5]
[I .0] [2.3]
.
0.23
K2] (d 1)
[16.1]
[64.6]
. . .
[672] [720]
[317]
[46] [250]
! 0.7 13.5
[24]
26
. . .
, Ki (mL/g werld ')
D i ( 2 - E t h y l h e x y l ) P h t h a l a t e (,:ont.)
Feeding
Test Conditions
.
.
Total
-
--
--
--
78 113 42
---
--
-
.
220 - 270
.
40'
114
BCF j
.
1.0 - 29.7
51.5 8.9 1.6
15 19 16
962 u 6,510 u
637
6 !3!
Parent BCF I (mL/g wet "]) (mL/g w e r ' )
Table 7. Summary o f Bioconcentration Studies for Phthalate Esters with Fish.
16
15
14 14 14
13 13 13
12,17 12,17 12,17 12,17 12,17
2
II
3
I
Ref.
¢>
25
12
26
Temp (°C)
FT
FT
S
Expos.
30 - 40
6.4
O. I - 1.0
.
[70]
.
.
.
.
K it (mL/g wetld I)
Diisodecyl Phthalate 2% bdw
No
Y
Cone. ~g/L)
Exposure
Diisooctyl Phthalate
Feeding
Test Conditions
.
.
.
.
.
K~ (d -I)
.
.
Total
Parent
.
.
.
207 .
<3.6 - < 14.4
92
BCF ~ BCF ~ (mL/g wet ~) (mL/g wet -~)
t6
3
7
Ref.
References: 1, Barrows et al. 1980; 2. Wofford et al. 1981; 3. Boese, 1984; 4. Heidolph and Gledhill, 1979; 5. Carr et al. 1992; 6. Call et al. 1983; 7. Sanbom et al. 1975; 8. Metcalfet al. 1973; 9. Mayer, 1976; 10. Macek et al. 1979; 11. Freitag et al. 1985; 12. Karara and Hayton, 1989; 13. Mehrle and Mayer, 1976; 14. Tart et al. 1990; 15. Park et al. 1990; 16. CITI, 1992; 17. Karara and Hayton, 1984; 18. Scholtz and Diefenbach, 1996.
S = Static; FT = Flow Through; NR = Not Reported; % bdw = Percent of Body Weight Fed Daily; ad. lib. = ad libetum; [ ] Denotes values that are based on parent compound; :[: = Nominal concentration at start of test.
Footnotes Values reported are based on total radioactivity determined in whole body and water and do not distinguish parent compound from metabolites unless otherwise indicated. 2 BCF after 17 day exposure. 3 BCF for fillet. 4 BCF after 3 day exposure. s Calculated using fraction of total radioactivity reported as parent compound after 11 days. 6 Calculated from reported tissue concentration after 1 hour exposure. 7 Calculated from data obtained during uptake phase.. s Calculated from data obtained during deputation phase. 9 Calculated by multiplying BCF for parent compound by the i st order elimination rate of parent compound from deputation phase. ~0 Includes dietary exposure in which daphnids containing 28 IJg/g DEHP were fed to fish at a 5% bdw ration. '~ BCF test performed under elevated temperatures (29-35 C).
(Carp)
Cyprinus carpio 30
38.5 + 2.7
Parophrys vetulus
(English Sole)
NR ( - 0.1)
Weight (g wet)
(Mosquito Fish)
Gambusia sp.
Organism
Table 7. Summary of Bioconcentration Studies for Phthalate Esters with Fish.
oo
Crustacea
Algae
Mollusca
lnsecta
30 21 21 21 2!
Artemia
Daphnia magna
Hexagenia bilineata
Ischnura verticalis
Chironom~
FT FT
FT
21 2! 21
21
Selenastrum
Daphnia magna Daphnia magna Daphnia magna
S
24
S
SR
SR
SR
SR
S
FT
Physa
phAmosz~s
! O-12
SR
21
Gammarus pulex
pseudolimnaeus
Gammarus
Palaemontes kadiakensis
Cmstacea
SR
S
HR 21
S
NR
Selenastrum
Algae
S
Expos.
30
Temp. (°C)
Test Conditions
Pseudomonas
Organism
Bacteria
Taxa
.
2700-13,900 0.08 0.08 0.1 0.18
No No No No No
0.08 0.21 0.21 0.21
No Yes Yes
Dihexyl P h t h a l a t e
1.4
NA
No
Butylbenzyl Phthalate
100~:
Yes
128 129 371
.
33
350
133
99
189
.
.
416
0.1
NO
248
0.08
.
No
2,000 10,000
NA NA
1,000 - 7,000
NA
.
.
.
.
.
.
.
.
.
.
.
.
.
KI I (mL g weted "l)
Di-n-Butyl P h t h a l a t e
Feeding
Exposure Cone. (/zg/L)
.
.
.
.
.
.
. 0.13 0.09 0.07
.
0.33
0.5
0.29
0.14
0.47
0.28
NC
.
.
.
K21 (d "l)
.
.
.
. 999 !,414 s 5,2699
100
700
4583
714
403
3453
1857
1485
7503.6
( m L g wet "l)
Total BCFI
Table 8. S u m m a r y o f Bioconcentration Studies for Phthalate Esters with Other Aquatic Organisms.
----
1,3752
--
--
--
--
--
--
__
--
--
1,324
5,4752'3
78 - ! 802'3
( m L g wet "N)
Parent BCFI
8 8 8
8
7
4
3
4
4
6
5
4
3
2
1
Ref.
26
26 20-25
Daphnia magna
Culexpipiens
Dedogonium
Chlorella
Crustacea
Insccta
Algae
26
26 26 24 15 15 1° I° 1° 18 ° 18 °
Elodea
Physa
Physa
Physa
Mytilus edilus Mytilus edilus
Pitar morrhuana Pitar morrhuana Pitar morrhuana Pitar morrhuana Pitar morrhuana
Aquatic Maerophytes
Mollusca
26
26
26
ehysa
Algae
FT FT FT FT FT
FT FT
S
S
S
S
S
S
S
S
S
S
Expos.
Test Conditions Temp. (°C)
Mollusca
Organism
Dedogonium
Taxa
6.4
6.4
6.4
6.4
Yes Yes Yes Yes Yes
Yes Yes
No
Yes
No
NA
NA
NA
0.6 5.9 59 0.2 1.2
4 42
0.35
19
100
100
50
19
K~~
473 525
76
.
73
.
.
.
.
.
.
.
.
.
.
.
.
.
.
(mL g wet-ld")
Di (2-Ethylhexyl) Phthalate
Yes
Yes
Yes
NA
Cone. (,ug/L)
Di-n-Octyl Phthalate
Feeding
Exposure
.
.
.
.
.
.°
-.
0.20 0.20
0.29
0.09
--
(d-I)
K21
2,366 2,627
264
1,050
814
987
1,338
1,429
699
8,412
(mL g wet"l)
Total BCFt
Parent BCF I
1,3805 1,047 436 1,122 363
377
811
2754
5,4002
946
91
26
133
278
(mL g wet-I)
Table 8. Summary of Bioconeentration Studies for Phthalate Esters with Other Aquatic Organisms. (cont.)
13 13 13 13 13
12 12
7
10
10
10
II
10
9
9
9
9
Ref.
"-,.I
Cmstacea
Mollusca (cont.)
Taxa
D
10- 12 25 25 NR
Gammaruspulex
Asellus brevieaudus
26
22 20 20 20
Daphnia magna
Daphnia magna
Daphnia magna
20
21
Daphnia magna
vannamei
Panaeus
SR SR
21 2!
Gammarus pseudolimnaeus
SR SR SR SR
S
S
SR
FT
SR SR
FT FT FT FT FT FT
l* 1* 1* 18* 18 * 18*
Nucula annulata Nucula annulata Nucula annulata Nucula annulata Nucula annulata Nucula annulata
FT F'I" FT FT FT FT
Expos.
1* 1* 1* ! 8* 18" ! 8"
Temp. (*C)
Feeding
Cone. 0zg/L)
Exposure
Yes Yes Yes Yes
No
No
No
Yes
No No
Yes
No No
Yes Yes Yes Yes Yes Yes
Yes Yes Yes Yes Yes Yes
2.2 7.3 25.3 85.6
5.5
100:[:
0.3
60 - 6,000"
1.9 62.3
100
0.10 62.8
0.6 5.9 59 0.2 1.2 15.5
0.6 5.9 59 0.2 1.2 ! 5.5
. . . -
.
NC
61
--
37 NC
.
822 8.3
. . . . . .
. . . . . .
. . .
.
.
. . . . . .
. . . .
. .
. . .
.
.
. . . . . .
. . . .
.
(mL g wetld a)
K~t
Di (2-Ethylhexyl) Phthalate (cont.)
Test Conditions
Mulinia lateralis Mulinia lateralis Mulinia lateralis Mulinia lateralis Mulinia lateralis Mulinia lateralis
Organism
. . .
.
.
.
.
. . . . . .
. . . .
.
--
NC
NC
0.81
3.4 NC
0.21 0.10
. . . . . .
. . . .
.
K2J (da)
.
. . . . . .
. . . .
.
166 140 261 268
518-'
1834
41612
--
113 383'~°
14907
3916 83
Total BCF I (mL g wet"l)
-
-
179
1,995
1,820
174 100 1,9O5
16 16 16 16
15
10
4
14
3 3
5
3
4
13 13 13 13 13 13
! 26 s
933 1,175 3,31 I
3,890
13 13 13 13 13 13
Ref.
2,455 s 3,090
Paint BCF I (mL g wet "m)
Table 8. Summary o f Bioconeentration Studies for Phthalate Esters with Other Aquatic Organisms. (cont.)
21
21 22 26 5°
26.5 26.5
15 15 20 20 20 20
Hexagenia bilineata
Chironomus plumosus
Culex pipiens
Rana arualis
Diplora strigosa
Arca zebra
Mytilus edilus Mytilus edilus
Daphnia magna Daphnia magna Daphnia magna Daphnia magna
Insecta
Amphibia
Coelenterata
Mollusea
Mollusea
Crustacea
18 o 18 °
1o 1° 1°
NR
incisa incisa incisa incisa incisa incisa
SR SR SR SR
FT FT
S
S
S
S
SR SR
SR
D
FT FT FT FT FT FT
Expos.
Feeding
Exposure Conc. (/zg/L)
Yes Yes Yes Yes
Yes Yes
No
No
2.9 9.6 32.6 100
4.4 41.7
Diisodeeyl Phthalate
61
61
Phthalate
0.9 - 187
19
0.3 0.2
0.1
Diisononyl
Yes
Yes
No NR
No
10.05
5.9 59 0.2 1.2 15.5
Yes Yes Yes Yes Yes Yes
0.6
Yes
. . . .
795 539
848
--
.
.
568 81.6
128
--
. . . . .
.
. . . .
. .
. . . . .
.
. . . .
. .
. . . . .
.
K~~ (mL g wet-ld -I)
Di (2-Ethylhexyl) Phthalate (cont.)
Test Conditions Temp. (°C)
Lumbriculus terrestris
Nepthys Nepthys Nepthys Nepthys Nepthys Nepthys
Organism
Annelida
Polychaeta
Taxa
. . . .
.
. . . . .
.
.
0.20 0.18
0.46
0.02
.
1.8 0.2
0.14
0.04
. . . . .
.
K2 t (d "l)
.
. . . . .
.
100 147 128 90
3,977 2,998
1,844
--
1,892
315 408
915
.
.
.
Total BCFI (mL g wet -I)
- .
- .
. -
16 16 16 16
12 12
20
20
19
450 - 76013
4 18 10
- -
4
17
13 13 13 13 13
13
Ref.
1,892
.
47 30 I,I 22 724 549
595
Parent BCFI (mL g wet "l)
Table 8. Summary o f Bioconeentmtion Studies for Phthalatc Esters with Other Aquatic Organisms. (cont.)
I O
References: 1. Wang and Gradie, 1994; 2. Casserly et al. 1983; 3. Sanders et al. 1973; 4. Mayer and Sanders, 1973; ft. Thuren and Woin, 1988; 6. Hudson et al. 1981; 7. Parkerton, 1993; 8. Gloss and Biddinger, 1985; 9. Sanborn et al. 1975; 10. Metcalf et al. 1973; I 1. Geyer et al. 1981; 12. Brown and Thompson, 1982b; 13. Perez et al. 1985; 14. Hobson et al. 1984; 15. Macek et al. 1979; 16. Brown and Thompson, 1982a; 17. Albro et al. 1993; 18. Streufert et al. 1980; 19. Larsson and Thmen, 1987; 2O. Solbakken et al. 1985.
NR = not reported; NC ---not calculable due to erratic nature of reported tissue concentration during uptake phase; FT = flow through exposure; SR = static renewal exposure; S = static exposure; D = dietary exposure; 1: = nominal concentration at start o f test
Table 8. Summary of Bioconcentration Studies for Phthalate Esters with Other Aquatic Organisms. (cont.) Footnotes: = Values reported are based on total radioactivity determined in whole body and water and do not distinguish parent compound from metabolites unless= otherwise indicated. 2 BCF after I day. 3 Literature value converted from day to wet weight basis by assuming a dry/wet weight ratio of 0. ! 5. 4 BCF after 2 days. 5 BCF after 30 days in marine microcosm; calculated values include potential dietary exposure. 6 BCF after 3 days; steady-state not reached. BCF after 10 days based on the sum o f reported absorbed and adsorbed tissue concentrations. s Includes dietary exposure in which algae containing 0. I 1 ttg/g were fed to daphnia at a 1% bdw ratio. 9 Includes potential dietary exposure in which uncontaminated algae were fed to daphnia at a I% bdw ratio. io BCF after 21 days. is Concentration (ppm) administered in diet at 1% bdw ration for 4 days. 12 BCF after 21 days. J3 BCF after 60 days.
t.J
..,O
714 ~5th increased exposure concentration (Mayer, 1976). Boese (1984) hypothesized that this inverse relationship was likely due to test concentrations exceeding water solubility since the bioavailability of colloidal/particulate chemical species are expected to be reduced. Elevated temperature appears to reduce the elimination rate and significantly increase the overall BCF (Karara and Hayton, 1989). This observation may be due to the inhibition of temperature-sensitive enzymes that mediate biotransformation of phthalate esters. Application of chemical inhibitors that block enzymatic hydrolysis of diesters has also been shown to enhance phthalate ester bioconcentration in fish (Melancon et al., 1977; Karara and Haytorl, 1988).
In an attempt to compare phthalate ester toxicokinetic data for fish with results for other chemical classes, bioenergetics models were used to estimate K~ and K,_. K~ was calculated as described by Thomalm (1989): R
K] :
(3)
where: KI
= Uptake clearance [mL g wet t day ~]
R
= Respiration rate [mg 02 g wet" day ~]
E
= Efficiency of chemical uptake relative to that of oxygen [unitless]
Cox
= Dissolved oxygen concentration [mg 02 mL ~]
Respiration was estimated from fish weight (W, g wet) and experimental temperature (T, °C) using the empirical equation presented by Thurston and Goerke (1987): R = 1.32W -°]gs T °'29 exp °'°3s r
(4)
The ratio of the chemical to oxygen uptake efficiency (E) for fish was estimated based on the literature review by Parkerton (1993) for nonionic organic chemicals: For Log K.,, < 4; E = 1 - 0.3 (4 - Log Ko~,)
(5)
For Log Ko. ~ 4 but ~ 6, E :
1
(6)
and for Log Ko~ > 6; E
= 1 - 0.3
(Log
K,,,
-
6)
(7)
715 For these calculations, the dissolved oxygen concentration at saturation for the prescribed experimental conditions was assumed. For several studies included in Table 7, fish weights were not reported so reasonable values we're assumed based on literature data. The elimination rate was estimated according to Thomann (1989): Ki = - -
(S)
where f,pis the lipid fraction of the fish.A typical ft~pof 0.05 was selected for calculations. A comparison of predicted and observed phthalate ester toxicokinetic parameters for fish is illustratedin Figure 5. The upper part of this figure indicates that the predicted uptake clearances systematically overestimate observed values. This bias may be attributed to presystemic metabolism of phthalate esters by esterases associated with the fish gill (Barren et al., 198g). As previously mentioned, a second explanation contributing to the lower observed KI values may be bioavailabilitylimitationsassociated with performing bioconcentration testsat concentrations in excess of water solubility.However, this rationale would only apply to the higher molecular weight phthalate esters.
Examination of the lower panel in Figure 5 shows a similar comparison for elimination rates. The predicted elimination rates overestimate data in the low Ko~ range (i.e., DMP, DEP) which is likely due to limits placed on chemical diffusion rates from aqueous fish compartments. In contrast, predicted elimination rates grossly underestimate observed rates for higher molecular weight phthalate esters since metabolism is ignored. The difference between the predicted and observed elimination rates can in fact be used to provide quantitative estimates of biotransformation rates (Sijm, 1992; DeWolfe et al., 1992).
Experimental fish BCFs from Table 7 are summarized graphically in Figure 6 as a function of log K~. The observed BCFs do not appear to be well correlated to Kow and tend to cluster around a value of 100. Comparison with predicted BCFs calculated from equations (3) and (8), and denoted by the solid line, indicates the model overestimates observed values by
approximately three orders of magnitude for the higher molecular weight phtha[atc esters (Figure 6). Again, these results are attributedto metabolism (Hogan, 1977; Metcalfet al.,1973; Carl 1992; Ban'on et al.,1995). In con~st, BCFs for the more water soluble phthalatc esters arc underestimated by about an order of magnitude due to the erroneously high elimination mtc that is predicted from equation (5). Data for other aquatic organisms also indicate BCFs are lower than expected based on simple lipid-water partitioning and do not correlate with phthalate ester hydrophobicity (Table 8).
For studies in which BCFs are reported in terms of both total '4C and parent phthalate ester, results are highly variable. BCFs for parent phth~ate esters range from less than a factor of two to as much as a factor of fiRy lower than BCFs
det=nnlned from totalradioactivity(Tables 7 and g). To provide a furthersynthesis of available BCF data,the average BCF for various taxa were calculated for the two most commonly studied phthalate eaters, DnBP and DEHP (Table 9). The average taxon BCF was determined by first calculating the mean BCF for each species within a taxa and then computing
716
1000
(a) . . . . . . . . . . . . . . . .
•
0-.
o~"o
u
~
°°°
~a
°
J
'
°
' " ~
•
v/~:l
/
"
"°" °o
°
0 DMP DBP t '~vBBP
I
100
0 ~P ~DIOP
0
10
i
,
i
,
,,l,I
,
a
i
10
. . . . .
I
. . . . . . .
100
1000
Observed Uptake Olearanoe (mL g"a"~)
100.00000 10,00000
li
1.00000
#. 0,10000
DEHP D~P
0.01000 0.0010O
.,t
8~
0.00010 0.0O001 0,001
ee °
%
. . . . . . . .
i
0,010
. . . . . . . .
i
0.100
. . . . . . . .
i
. . . . . .
1.000
10.000
Obeerv~J Bl~natlon Rate (d"~)
Figure 5. Comparison of predicted and observed uptake clearances (a) and elimination rates Co) of phthalate esters in fish based on total radioactivity (open'symbols) or parent compound (filled symbols).
717
7
I
I
I
I
I
I
0 DMP ADi~ oDBP vBBP ODEHP ODIOP
6
4 8
1 I
I
I
i
I
I
2
8
4
5
6
7
8
Log Kow
Figure 6. Comparison of measured fish bioconcentration factors based on total radioactivity (open symbols) or parent compound (filled symbols) with model predictions (Hne) as a function of the octanol-water partition coeffident (Log K~).
718 Table 9. Summary of Mean Bioconcentration Factors for DBP and DEHP in Different "faxa.
Phthalate
Taxon
# Species
Mean BCF (ml g wet 1)
Di-n-Butyl
Algae
1
3399
Phthalate
Crustaceans
3
662 ± 229
Insects
3
624 ~- 144
Fish
1
167
Di (2-Ethylhexyl)
Algae
2
3173 + 3149
Phthalate
Molluscs
5
1469 ± 949
Crustacea
4
1164 + 1182
Insects
3
1058 ± 772
Polychaetes
1
422
Fish
5
280 ± 230
Amphibians
1
605
+ standard deviation of species mean BCFs
the mean of the species BCFs for that taxa. In cases where BCFs for parent phthalate ester were provided, these values were used in calculations. Otherwise, BCFs based on total ~4C can be used to provide conservative BCFs for parent phthalate ester. Results indicate mean BCFs are highest for algae and lowest for fish with invertebrates exhibiting intermediate values. These findings are consistent with studies by Wofford et al. (1981) who found that the extent of phthalate ester biotransformation increased as follows: molluscs < crustaceans < fish. It is interesting to note that the lowest elimination rate reported in Table 8 was observed for the marine coral,
Diplora strigosa, which is consistent with the more primitive
enzymatic system expected for coelenterates.
The specific metabolic pathway of phthalate esters in aquatic organisms has been investigated in a number of studies and reviews (Stalling et al., 1973; Johnson et al., 1977; Melancon, 1979; Lech and Melancon, 1980; Barton et al., 1995). The first step in the metabolism of phthalate esters is hydrolysis to the monoester and the corresponding alcohol. The alcohol is further metabolized via a carboxylic intermediate to acetate and carbon dioxide by [3-oxidation. The monoester can be: directly excreted; undergo Phase I reactions in which the alkyl chain is subjected to various oxidations via microsomal P450s and mitochondrial enzymes; or undergo Phase II reactions in which monoester is conjugated with glucuronide (Barton et al., 1995). Phthalate ester metabolism appears to depend upon both species and exposure route. Earthworms
719 lppear to be somewhat unique in lacking the capacity to oxidize monoester which may account for the relatively low ,-limination rate observed for this organism (Table 8).
Although numerous studies have investigated the bioconcentration of phthalate esters, comparatively little work has been reported on potential uptake through food. Dietary accumulation is usually described as a simple first order process (Bruggeman, 1981). At steady state, the concentration in a predator is given by:
c~,=Nc K2 prey
(9)
Where •
Cp,,,Lp,y = chemical concentration in predator or prey (diet) [~g g wet ~] I
= ingestion rate [g wet prey g wet predator'ld"~]
(X
= assimilation efficiency of dietary contaminant, and
K2
= the elimination rate defined previously.
Ma~k et al. (1979) examined dietary transfer to bluegills that were fed DFJ-IP contaminated daphnids while Hobson et ai. (1984) examined dietary uptake of DEHP by penseid shrimp fed a synthetic diet. Application of equation (9) to results based on total ~4C measurements obtained in these studies yield assimilation efficiencies of ca. 0.25-0.30. Assimilation efficiencies for pment DEHP may be lower as a consequence of metabolism in the gut. Gloss and Biddinger (1985) studied the importance of dietary transfer in daplmids fed DHP contaminated algae, but equivocal results preclude estimation of reliable assimilation efficiencies. In one set of experiments, accum,laHon from food+weter was 39% higher than water only exposures. However, inconsistent results were obtained in food-only experiments since no accumulation was observed. Perez et al. (1985) suggested that dietary exposure could account for seasonal ~
in the accumulation of DEHP by
marine biota in microcosm studies. Bivalve filter feeders had tenfold higher tissue concentrations in winter compared to summer experiments despite only a threefold increase in average DEHP water concentrations. It was suggested that the higher tissue concentrations could be explained by higher algal densities and subseqtamt ingestion that ocom'red during the winter or that lower winter temperatures decreased enzymatic activity. In contrast, benthic species had higher tissue concentrations in the summer that were attributed to elevated sediment ingestion rates associated with warmer ~
.
The relative importance of dietary and aqueous exposure can be determined by comparing the ratio of dietary and aqueous uptake:
¢~'~ = --~-~BAF
K I C,w ,
K,
(10)
720 To illustrate the relative importance of these mutes for DEHP, representative values for the parameters included in equation (10) can be selected for a 0.5 kilogram fish that feeds on invertebrates. An assimilation efficiency of 0.3 and an uptake clearance of 4 ml g wet1 day~ were selected based on laboratory studies (Macek et al., 1979; Tarr et al., 1990). A prey BAF of 2000 ml g wet"~was assumed based on the microcosm studies by Perez et al. (1985) while an ingestion rate of 0.01 g wet prey g wet predator "t day "1was obtained from the allometric equations suggested by Thomann (1989). Substituting these estimates into equation (10) yields a value of 1.5 indicating that as much as 60% of the DEHP exposure could be derived from the diet. In contrast, if the diet was comprised of smaller fish rather than invertebrates the prey BAF would be an order of magnitude lower and the dietary route would account for less than 15% of the DEHP accumulated. These results differ from estimates obtained for persistent organochlorine compounds with a Kow similar to DEHP which indicate more than 95% of the contaminant accumulation is due to the dietary route (Thomann, 1989). Consequently, increased metabolism and reduced prey concentrations between trophic levels prevents biomagnification of phthalate esters. In fact, phthalate ester concentrations are expected to decrease as one proceeds up the foodchain. This inverse pattern has been observed for other classes of metabolizable organic chemicals such as polyaromatic hydrocarbons (McElroy et al., 1989).
Bioaccumulation of sediment-associated phthalate esters by dragonfly and chironomid larvae has been investigated (Woin and Larsson, 1987; Brown et al., 1996). Assuming a typical lipid content of 1% wet weight for benthic invertebrates, biota to sediment accumulation factors (BSAFs) can be calculated. The BSAF expresses the ratio of the lipid-normalized biota concentration at steady-state to the corresponding organic carbon- normalized sediment concentration to which test organisms are exposed. The BSAF for DEHP in dragonflies is approximately 0.1 while the BSAF for both DEHP and DIDP in chironomids is about 0.5. These values likely overestimate BSAFs for parent compound since calculations are based on total 14C.Nevertheless, these BSAFs are lower than predicted from equilibrium partitioning theory (Lake et al., 1990) further supporting the importance of phthalate ester metabolism.
Imlma~_Qzumi.tz~ Bioaccumulation of phthalates by terrestrial plants has been the subject of several studies. Plant cuticle lipid-water partition coefficients were determined for three plants with a range of organic chemicals including DEHP and were found to be approximately equal to K ~ (Kerler and Schonherr, 1988). However, these studies employed radiolabeled compounds and did not determine if parent compound was accumulated. Shea et al. (1982) reported the uptake of DnBP contaminated soil by com. Plant to soil bioconcentration factors, defined as the ratio of wet weight concentrations in plant and soil, respectively, were less than 0.002 indicating the limited bioaccumulation potential of DnBP. Kirchmann and Tengsved (1991) investigated the uptake of DBP and DEHP in barley grown on soil fertilized with sludge containing these compounds. Application of sludge containing 37 mg/kg of DBP and 116 rdg/kg of DEHP to field plots at a rate of 5 tom/ha resulted in concentrations in barley at harvest of 0.116 mg/kg and 0.530 mg/kg dry, respectively. Concentrations
721 in plants harvested from control plots receiving no sludge addition were < 2 and ca. 5 time lower for DBP and DEHP, respectively. The authors concluded that uptake percentages of added phthalates by grains was very low amounting to only 0.1-0.2 % of the amount added to the soil. Overcash et al. (1986) investigated the uptake of t4C-DEP, DnBP, DnOP and DEHP from soil by corn, soybean, wheat and rescue. Plant species were grown to maturity in greenhouse pots in soil containingthree different treatment concentrations of each phthalate ranging from 0.02 to 4 mg/kg. Total ~4Cin plants at harvest indicated limited plant uptake with plant bioconcentration factors, expressed as the ratio of the plant concentration (wet weight) to the initial soil concentration (dry weight), ranging from <0.001 to 1.0 with most values < 0.1 and typical values <0.01 (Overcash et al., 1986). Since these values are based on total '4C, these values overestimate the true extent to which parent phthalates are bioconcentrated as supported by the findings of additional studies described below.
Uptake of )4C-DEHP by four species of plants (i.e., lettuce, carrot, chili peppers and tall rescue) grown on soil amended with DEHP contaminated municipal sludge was investigated by Aranda et al. (1989). Soil DEI-IP concentrations ranged from 2.6 to 14.1 mg/kg dry in these studies. Since parent DEHP was not detected in any of the plants grown on these soils, the authors concluded that plant uptake of DEHP was of minor importance and would not limit sludge application to soils used to grow the crops studied. These findings are consistent with laboratory and field studies reported by Schmitzer et al. (1988) who found that parent DEHP was not detected in barley and potatoes grown in soils contaminated with DEHP at concentrations ranging from 0.2 to 3.3 mg/kg. The low levels of total ~'C detected in plants from both studies were attributed to uptake of ~4COz produced from the biodegradation of ~4C DEHP during the course of the study. This explanation is further supported by the results of Kloskowski et al. (1981) who observed the mineralization of 14C DEI-IP in soil-plant systems. Schmitzer et al. (1988) cites earlier studies by Kato et al. (1980) who also reported DEHP was not accumulated by vegetables grown in DEHP contaminated soils. Foliar uptake of DBP and DEHP in air by plants also appears to be limited (Lokke and Bro-Ramussen, 1981, and Lokke and Rasmussen, 1983).
Data on the bioaccumulation of phthalates in terrestrial animals is limited but, like aquatic foodchalns, trophic transfer is expected to be of low concern as a result of extensive biotransformation. In one study, a simple detrital foodchaln consisting of woodlice feeding on fallen,4cer leaves contaminated with DEHP was used to investigate bioaccumulation potential in the tenestrial environment. Results indicated no DEHP bioconcenUation in adult woodlice or offspring during the 6 month microcosm experiment and showed that DEHP is degraded effectively in this terrestrial foodchain (Lokke, 1988). Several studies have investigated the bioaccumulation of selected phthalate esters in avian species. Belise et al. (1975) found that DBP and DEHP did not accumulate in tissues of mallard ducks above the limit of detection (0.15 mg/kg) after being fed diets comaining 10 mg/kg of either chemical for 5 months. Captive starlings were fed diets containing 25 and 250 mg/kg of DHP or DEHP for 30 days by O'Shea and Stafford (1980). Residue analyses at test termination indicated that DHP was not detected ( < 1 mg/kg) in any of the eight carcasses examined from either dietary dose. Five out of the eight birds contained detectable levels of DEHP. The average residue in samples with detectable concentrations was 1.8 mg/k8. In
722 another study, Ishiuda et al. (1982) fed white leghorn hens contaminated diets containing 50,000 and 100,000 mg/kg of DEHP for 230 days.
Various tissues were analyzed in two hens from each treatment at test termination. DEHP
concentrations in various muscles tissues ranged from 1.8-9.4 and 2.6-16.4 mg/kg for the 50,000 and 100,000 mg/kg dietary concentrations, respectively. In a subsequent 25 day feeding study in which hens were fed diets containing 2000 mg/kg of DEHP, muscle tissue concentrations ranged from <0.5 to 1.5 mg/kg. In both experiments, higher concentrations were observed in adipose tissues and feathers than muscle tissues. The elevated concentrations in feathers likely resulted from external contamination of the birds from the diet during feeding. Based on the results of the~ studies it is apparent that phthalate esters exhibit a very low bioaccumulation potential in avian wildlife.
Predictions that rely on simple correlations with chemical hydrophobicity for estimating bioaccumulation in agricultural foodchains such as proposed by Travis and Arms (1988) are inappropriate if applied to phthaiates since these models arc based on data for persistent chemicals. In a recent critique of the application of this model for assessing potential indirect human exposure to DEHP, Lowenbach et al. (1995) suggested an alternative approach to conservatively estimate bioaecumulation in meat and dairy products derived from contaminated areas. This approach relies upon an empirical correlation derived between bioconcentration in rodents and bioconcentration in earle for a large set of chemicals. Based on a measured rodent BCF for DEHP, the cattle BCF was estimated to be 0.008. This BCF represents the ratio of DEI-IP concentrations in cattle to that in the cattle's food. This estimate is three orders of magnitude lower than obtained using the Travis and Arms (1988) model which fails to take into account the importance of metabolism in limiting bioaeeumulation. Consequently, use of such simple models should be avoided to prevent exaggerating phthalate ester exposures via the agricultural foodchain.
SUMMARY 1. A large body of existing data on phthalate ester physieochemieal properties is available. This chemical class exhibits an eight order of magnitude increase in octanol-water partition coefficients (K,~) and four order of magnitude decrease in vapor la'essure (VP) as alkyl chain length increases fi-om I to 13 carbons. Phthalate esters span a twelve order of magnitude change in aqueous solubility over the range of alkyl chain lengths. A critical review of water solubility measurements for higher molecular weight phthalate esters (i.e. alkyl chains ~ 6 carbons) reveals that most published values significantly exceed true values due to exper~ental difficulties associated with solubility determinations for these hydrophobic organic liquids. The higher molecular weight phthaiate esters have extremely low water solubilities.
2. Laboratory and field studies show that partitioning to suspended solids, soils, sediments and aerosols increases as K ~ increases and VP decreases.
723 3. Atmospheric fate processes including photooxidation, washout, and vapor-aerosol partitioning were reviewed. Photodegradation via free radical attack is expected to be the dominant degradation pathway in the atmosphere with predicted half-lives of ca. 1 day for most of the phthalate esters investigated.
4. Biodegradation is expected to be the dominant degradation removal mechanism for phthalate esters entering sewage treatment plants, or released to surface waters, soils and sediments. Numerous studies indicate that phthalate esters are degraded by a wide range of bacteria and actinomycetes under both aerobic and anaerobic conditions. Standardized biodegradation tests with sewage sludge inocula showed phthalate esters undergo ~ 50% ultimate degradation within 28 days with many studies showing ~ 90% removal. Primary degradation half-lives in surface and marine waters range from <1 day to 2 weeks and in soils from <1 week to several months. Longer half-lives may occur in anaerobic, nutrient poor, or cold environments.
5. Numerous experiments have shown that the bioaccumulationof phthalate esters in the aquatic and terrestrial foodchain is limited by biotransformation, which progressively increases with trophic level. In general, bioconcentration factors decrease for organisms that possess more advanced metabolic capabilities. Overall, BCFs span only about three orders of magnitude across phthalate esters. These data indicate that current bioaccumulationmodels developed for persistent organic chemicals that ignore biotransformation grossly exaggerate the bioaccumulation potential of phthalate esters.
6. The results of this study provide the technical basis for future efforts aimed at quantifying multimedia environmental exposure to phthalate esters.
724
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Pergamon PII: S0045-6535(97)00195-1
0 1997 Elsevier Science Ltd All rights reserved. Printed in Great Britain 0045-6535/97 $17.00+0.00
THE ENVIRONMENTAL FATE OF PHTHALATE ESTERS: A LITERATURE REVIEW
Charles A. Staples ~, Dennis R. Peterson2, Thomas F. Parkerton2 and William J. Adams3
Assessment Technologies, Inc., 10201 Lee Highway, Suite 305, Falrfax, VA 22030 USA I Exxon Biomedical Sciences, Inc., Mettlers Road, CN 2350, East Millstone, NJ 08875 USA2 Kennecott Utah Copper Corp., Magna, UT 84044 USA 3 (Received in Germany 4 February 1997; accepted 5 March 1997)
ABSTRACT A comprehensive and critical review was performed on the environmentalfate of eighieen commercial phth~ate esters with alkyl chains ranging from 1 to 13 carbons. A synthesis of the extensive literature data on physicoehemical properties, partitioning behavior, abiotic and biotic transformations and bioaccumulationprocesses of these chemicals is presented. This chemical class exhibits an eight order of magnitude increase in cctanol-water partition coefficients (K~) and a four order of magnitude decrease in vapor pressure (VP) as alkyl chain length increases from 1 to 13 carbons. A critical review of water solubility measurements for higher molecular weight phtbadate esters (i.e. alkyl chains ~ 6 carbons) reveals that most published values exceed true water solubilities due to experimental difficulties associated with solubility determinations for these hydrophobic organic liquids. Laboratory and field studies show that partitioning to suspended solids, soils, sediments and aerosols increase as K ~ increases and VP decreases. Photodegradation via free radical attack is expected to be the dominant degradation pathway in the atmosphere with predicted half-lives of ca. 1 day for most of the phtl~ate esters investigated. Numerous studies indicate that phfludate esters are degraded by a wide range of bacteria and aetinomyeetes under both aerobic and anaerobic conditions. Standardized aerobic biodegradation tests with sewage sludge inocula show that phthalate esters undergo ~ 50% ultimate degradation within 28 days. Biodegradation is expected tobe the dominant loss mechanism in surface waters, soils and sediments. Primary degradation half-lives in surface and marine waters range from <1 day to 2 weeks and in soils from <1 week to several months. Longer half-lives may occur in anaerobic, oligotrophie, or cold environments. Numerous experiments have shown that the bioaccnmulationof phthalate esters in the aquatic and terrestrial foodchaln is limited by biotransformation, which increases with increasing trophic level. Consequently, models that ignore biotransformation grossly exaggerate bioaccumulation potential of higher molecular weight phthalate esters. This review provides the logical first step in elucidating multimedia exposure to phthalate esters. © 1997 Elsevier Science Ltd 667
668 INTRODUCTION ~01amaercial Uses and Products Phthalate esters are widely used industrial chemicals (ECETOC, 1988; Giam et al., 1984; Graham, 1973), serving as important additives which impart flexibility in polyvinylchioride (PVC) resins. Phthalate esters are also used to varying degrees in other resins such as polyvinyl acetates, cellulosics, and polyurethanes. The stability, fluidity and low volatility of higher molecular weight phthalate esters make them highly suitable as plasticizers.
Most of the mid- to high- molecular weight phthalate esters are used in the manufacturing of PVC, while di-n-butyl phthalate is used in epoxy resins and cellulose esters and specialized adhesive formulations. Dimethyl and diethyl phthalate esters are typically used in cellulose ester-based plastics, such as cellulose acetate and butyrate. Plasticizers are used in building materials, home furnishings, transportation, clothing, and to a limited extent in food (packaging) and medical products. Release of phthalate esters into the environment during manufacture, use, and disposal has recently been reviewed (Warns, 1987; Cadogan et al. 1993). The intent of this review is to evaluate the extensive literature on the physicochemical properties and the abiotic/biotic processes that dictate the ultimate fate and distribution of phthalate esters in the environment.
The eighteen commercially important phthalate esters reviewed are listed in Table 1, where the abbreviations used for the various phthalate esters are presented. In general, the abbreviation indicates when the phthalate ester is a mixture of branched or linear isomers (i.e., DnOP for linear di-n-octyl phthalate and DIOP for isomeric diisooctyl phthalate). Butylbenzyl phthalate (BBP), is non-symmetric, having both butyl and benzyl alcohol esters groups. With the exception of di-2(ethylhexyl) phthalate, which is a pure isomer, higher molecular weight phthalates (i.e. alkyl chains >6 carbons) are mixtures based on the alcohols used for synthesis. For example, 711P is a mixture of di-heptyl, nonyl, and andecyl phthalates. Likewise, 610P is a mixture of hexyl, octyl, and decyl isomers. Other products, such as diisononyl and diisodecyl phthalates are isomeric mixtures comprised principally of C9 or C10 alkyl chains. Considerable differences in structure are thus possible for certain commercial phthalate esters.
PHYSICOCHEMICAL PROPERTIES
Review of phthalate ester physical properties is important to help understand behavior and fate in the environment. A number of the important physicochemical properties of phthalate esters are reviewed below.
Common commercial phthalate esters are liquids at ambient temperatures (Table 2). Nearly all have melting points (or pour points) below -25 °C. The exceptions are DMP, DUP, and 610P with melting points of +5.5 °C, -9 °C, and -4 ° C,
669 Table 1. Eighteen Commercial Phthalate Esters Reviewed. Abbreviation
Phthalate Ester
CAS No.
DMP
Dimethyl Phthalate
131-11-3
DEP
Diethyl Phthalate
84-66-2
DAP
Diallyl Phthalate
131-17-9
DPP
Dipropyl Phthalate
131-16-8
DnBP
Di-n-Butyl Phthalate
84-74-2
DIBP
Diisobutyl Phthalate
84-69-5
BBP
Butylbenzyl Phthalate
85-68-7
DHP
Dihexyl Phthalate
84-75-3; 68515-50-4
DnOP
Di-n-Octyl Phthalate
117-84-0
BOP
Butyl 2-Ethylhexyl Phthalate
85-69-8
610P
Di (n-Hexyl, n-Octyl, n-Decyl) Phthalate
25724-58-7; 68515-5 I-5
DEHP
Di (2-Ethylhexyl) Phthalate
117-81-7
DIOP
Diisooctyl Phthalate
27554-26-3
DINP
Diisononyl Phthalate
28553-12-0; 68515-48-0
DIDP
Diisodecyl Phthalate
26761-40-0; 68515-49-1
D711P
Di (Heptyl, Nonyl, Undecyl) Phthalate
3648-20-2; 68515-44-6 68515-45-7; 111381-89-6 111381-90-9; 111381-91-0
DUP
Diundecyl Phthalate
3648-20-2
DTDP
Ditridecyl Phthalate
119-06-2; 68515-47-9
respectively. Phthalate esters have boiling points varying from about 230°C to 486°C. Higher molecular weight phthalate esters have boiling points that must be determined at reduced pressure to prevent thermal decomposition. The low melting point and high boiling point of these phthalate esters contribute to their usefulness as plasticisers, heat transfer fluids, and carders.
Water Soinbilitv Water solubility is an extremely important property that influences the biodegradation and bioaceumulation potential of a chemical, as well as aquatic toxicity. Water solubility also is a determining factor controlling the environmental distribution of chemicals. Losses from wastewater treatment facilities, landfills and sludge-amended soils are partially a function of aqueous solubility.
4 4, 6 ~
CI2HI404
CI4HI404
Ci4His04
Ct6H2204
C16H2204
DEP
DAP
DPP
DnBP
DIBP
7, 9,11 11
C24H3804
C~H3004
CzsH4004
C24H3sO4
C24H3sO4
C26H~204
C28H4604
C26H4204
C30H~O4
C34H5804
DnOP
BOP
610P
DEHP
DIOP
DINP
DIDP
D711P
DUP
DTDP
530.8 {506.8-544.8
447.7 {432.7-474.7
418.6 {362.6-474.7
446.7 {432.7-446.7
418.6 {418.6-432.6
390.6 {376.6-390.6
390.6
404.6 {334-447}
334.4 {278.4-390.6
390.6
334.4
312.4
278.4
278.4
250.3
246.2
222.2
194.2
Molecular Weight
1.111 1.011 0.978
-35 -27.4 -25
-37
-9
<-50
-46
-48
-46
-47
-4
0.953
0.96
0.97
0.961
0.97
0.986
0.986
0.97
--
1.050 c
-58
-37
1.042
1.118
-40
-35
1.192
Specific Gravity (20 °C)
5.5
Melting Point (°C)
Data from Howard et al. (1985); ttoward et al. (I 989); Sears and Touchette (1982); and manufacturer's unpublished data. For mixtures, the formula for the most representative isomer is provided. Aromatic ring { } molecular weight range of isomer At 15 °C
13
10
9
8
8
6, 8,10
6 c, 8
8
6
CI9H1oO4
C~H3oO4
BBP
DHP
4
3
3
2
1
CioHioO4
DMP
Alkyl Chain Length
Formula b
Compound
Table 2. Physical Properties of Eighteen Phthalate Esters?
.,.....I O
671 Precise water solubility measurements for chemicals with moderate to high water solubilities (> I mg/L) can be obtained easily with conventional methods e.g., shake-flask procedures. In contrast, water solubility measurements for more hydrophobie compounds can be confounded by a variety of experiment problems. First, the inability to separate colloidal emulsions of undissolved chemical from the aqueous phase may cause difficulties in determining accurate water solubilities (Yalkowsky and Banerjee, 1992). Another source of error of particular concern for phthalate esters is artifactual contaminationfrom laboratory plastics (Giam et al., 1984). Since higher molecular weight phthalate esters are slightly less dense liquids than water at ambient temperatures, these chemicals form surface films at the air-water interface. Consequently, errors may be introduced when withdrawing aqueous phase samples for analysis through surface films (Howard et al., 1985). Each of the above problems can lead to experimental artifacts that yield measured values that overestimate the true water solubility.
Recognizing the above limitations, reported aqueous solubilities ofphthalate esters are summarized in Table 3. Results show a declining trend in water solubility with increased carbon number of the alcohol moiety for the lower molecular weight phthalates through DHP. Moreover, independent experimental measurements are generally in good agreement and are consistent with theoretical predictions obtained by the SPARC (Karickhoffet al., 1991) as well as the EPIWIN (Meylan and Howard, 1995) structure-activity models. Thus, experimental measurements are believed to be reliable for lower molecular weight phthalates.
In contrast, experimental data for higher molecular weight phthalates often appear suspect. First, the expected decreasing trend in water solubility with increasing alkyl chain length is not apparent for these compounds (Howard et al., 1985). Furthermore, measurements obtained by independent investigators using different methods often vary by several orders of magnitude. For example, Howard et al. (1985) reported a value of 0.34 mg/L for DEHP using a shake-flask/centrifugation method. Similar results of 0.285 mg/L and 0.36 mg/L were reported by HoUifield (1979) and Defoe et al. (1990) respectively, using uephelometric techniques. In comparisojj, Leyder and Boulanger (1983) measured a water solubility for DEHP of 0.041 mg/L using a shake flask/centrifugation method. In seawater, Giam et al. (1980) and Howard et al. (1985) report DEHP solubilities of 1.16 and 0.160 rag/L, respectively, using conventional techniques. In contrast, Boese (1984) reported a value of 0.0006 mg/L in seawater using a generator column. This latter value is consistent with the predicted estimate of 0.0026 mg/L and 0.0011 mg/L obtained from the SPARC and EPIWIN models, respectively (Long, 1995; Meylan and Howard, 1995). Early studies reported that the water solubility of DINP and DIDP ranged from 0.2 - 1.19 mg/L. More recent experiments that employed a slow-stir shake flask method and included precautions to prevent sampling of ulidissolved test material indicate values of 0.0006 and <0.00013 for DINP and DIDP, respectively (Exxon Biomedical Sciences, Inc., 1996a). These measurements are in much closer agreement to theoretical predictions based on structure activity relationships (Table 3).
672 T a b l e 3. S e l e c t e d P a r a m e t e r s C o n t r o l l i n g t h e E n v i r o n m e n t a l Compound
............
DMP
1.46 (1) 1.47 (2) 1.48 b 1.53 (3) 1.56 (4) 1.60 (37) 1.61 (4)
Log K . - . . . . . . . . . . . . . . . . . . . . . . 1.61 (5) 1.62 (6) 1.62 (7) 1.66 (5) 1.66 d 1.74 (5) 1.90 (2)
2179 d 2810 (1) 3160 (2)' 3300 b 4000 (2)
D i s t r i b u t i o n o f P h t h a l a t e E s t e r s a.
AQ SOL (mg/L) . . . . . . . . . . . . . . . . . 4248 (4) 4290 (3) 4320 (18) 4200*
V P (mm Hg, ~25 °C) . . . .
1.65E-3 (2) 1.8E-3 (27) 5.4E-3 (38) 5.5E-3 c 2.9E-2 d 2.0E-3"
1.61' DEP
2.21 2.21 2.24 2.29 2.35
(6) (6) (2) (2) (3)
2.42 (37) 2.47 (8) 2.51 b 2.65 a 2.67 (4) 3.00 (9)
260 d 400 ~ 680 (19) 720 (20) 896 (18) 928 (3)
1080 (2) 4.8E-5 (27) 1160-1240 (21) 20°C 3.9E-4 (28) 1340-1400 (21) 30°C 6.1E-4 (28) 7028 (4) 8.1E-4' 1.2E-3 (35) 1100" 1.65E-3 (2) 5.0E-3 a
43 d
182 (3)*
3.7E-5 ¢ 1.6E-4 (35)*
108 (3)*
8.9E-5 c 1.04E-3 d
2.38* 1.0E-3* DAP
3.16 b 3.23 (3)*
3.36 d
DPP
3.27 (3)* 3.57 b
3.63 d
38 a 47 ~
DnBP
3.74 (2) 4.08 (10) 4,11 (11) 4.13 (6) 4.30 (12) 4.39 (10) 4.50 (37) 4.56 (10)
4,57 (3) 4.61 d 4.63 b 4.72 (8) 4.72 (13) 4.79 (2) 5.15 (31)
1.Y 11.2 (2) 3.2 (26) 13.0 (18) 3.25 (22)" 4.9 b 11.2' 8.%9.6 (13) 10.l (3) 10.5- I 1.1 (21 ) 20 °C 11.2-11.5 (21) 30°C
9.3E-6 (35) 1.7E-5 (32) 1.9E-5 (29) 2.08E-5 (30) 2.7E-5 (27) 3.5E-5 (34)
4.46 e
5.1 d 6.2 (23) 9,6 b
20.0(24) 20.3 (3)
1.8E-6 c 5.8E.-4 e
4.77 (11) 4.77 (14) 4.77 b 4.84 d 4.91 (3)
0.67 e 0.70 (23) 1.90 (20) 2.40 b 2.69 (2)
2.82 2.90 2.90 40.2
100b
3.6E-5 (36)
4.45* DIBP
4.11 (3) 4.31 b 4. I l *
BBP
3.57 3.97 4.05 4.11 4.73 4,75
(2) (6) (4) (11 ) (37) (4)
20.0 *
(3) (18) (5) (4)
2.7* 4.59*
DHP
8.7E-7 (35) 7,7E-6 e 8,6E-6 (14) 9.0E-6 (2) 9.1E-5' 5.0E-6"
5.65 (2)
6.67 ~
0.019 d
1.9E-8 ¢
5.93 (2) 6.57 e
6.82 (27)
0.049 b 0.24 (2)
1.8E-6 (34) 1.4E-5 (2) 1.2E-5 d
6.30* 0.05*
5.0E-6"
1.16E-3 d
4.1E-5 (30) 7.3E-5 (2) 1.2E-4 c 2.5E-4 d 2.7E-5"
673 Table 3. Selected Parameters Controlling the Environmental Distribution o f Phthaiate Esters (cont.) Compound DnOP
.........
Log K ,
5,22 (3) 7,06 (15) 8.06 (6)
8.10 (37) 8.30b ~,~4 d
AQ S O L (mg/L) . . . . . . .
VP (ram Hg) - ~
0.00046b 0.0005~ 0.001 (20)
0.02 (13) 0.02 (23) 3.00 (181 0.0005*
2.2E-7 (27) <3.4E-7 (35) 3.4E-7d
0.02d 0.11b*
1.1E-7 (35)* 2.4E-5d
1.2E-4~
8.06*
1.3E-5¢ 1.9E-4 (29) 1.0E-7*
BOP
6.28b* 6.5d
610P
5.9-8.6 (2) 7.25*
8.54d
0.0004d 0.05*
0.90 (2)
3.4E=7d 4.9E-6 (2)*
4.9E-4¢
DEHP
4.20-5.22 (16) 5.11 (ll) 7.06 (151 7.14±0,15 (16) 7.27 (37)
7.54b 7.80-8.90 (17) 7.86-8,15 (16) 7.94 (2) 8.06 (6)
0.0006 (25)' 0.0011 d 0.0026 b 0.041 (3) 0.16 (2)®
0.285 (23) 0.34(2) 0.40 (17) 0.40 (18) 1.16 (22)'
4.1E-8 4.5E-8 7.2E-8 9.8E-8 2.8E-7
<3.4E-7 (34) 3.8E-7 (27) 7.1E-7 (32) 6.4E-6 (2) 1.4E-6d
7.45 (15)
~,~9 d
0.27-0.36 (13)
1.2 (26")°
3.3E-7 (36)
7.45±0.06 (16)
7.50*
DIOP
8.00b* 8.39~
DINP
>8.0 (2)* 9.0b
0.003* 0.00024d 0.00081 b
0.09 (2)
1.4E-4~ 1.0E-7"
2.0E- I0c 3.4E-7 (35) 1.4E-6d
1.4E-6 (29) 5.6E-6 (21 1.0E-6*
0.0006 (39) 0.20 (21 <0.001 *
5.4E-7 (2)
2.2E-6 d 7.4E-6 ~
0.28 (24) 1.19 f21
5.1E-8 (36) 5.6E-8 (35)
0.001" 9.4d
(33) (12) (35) (I I) (29)
7.8E-5 b 2.3E-5 ~
<5.0E-7 *
DIDP
>8.0 (2)* I0,0b
<0.00013 (39)
<0.001*
<5.0E-7 (2)
<5,0E-7"
71 lP
6.0 (2) >8.0 (2)*
9.52d
1.7E-5d <1.0 (2)
<0.001"
5.0E-8d 2.8E-7 ( 3 5 )
<5.0E-7 (2"J <5.0E-7'
DUP
>8.0 (2)* 11.2b
11.5d
1.6E-7t' 4.2E-7b
1.10 (2)
5.3E-7 (35)
DTDP
>8.0 (2)* 13.1b
13.4d
1.5E-9d 4.2E-9b <0.3 (2)
10.3d
<0.001'
b ' d ° *=
0.34 (23) <0.001*
2.5E-11d
3.7E-8d
<5.0E-7' <5.0E-7 (2) <5.0E-7"
Includes measured and calculated values. Values calculated from the SPARC Structure Activity Program (Long, 19951. Calculated using the structure activity relationships in A S T E R (Russom et al., 1991). Calculated using the structure activity relationships in EPIWIN (Meylan and Howard, 1995). Solubility measurements made in seawater. R e c o m m e n d e d values based on available evidence.
References 1. Nielsen and Bundgaard, 1989; 2. Howard et al. 1985; 3. Leyder and Boulanger, 1983; 4. Veith et al. 1989; 5. Renberg et al. 1985; 6. McDuffie, 1981; 7. _P~dsforth, 1986; 8. Hansch and Leo, 1987; 9. DeKoch and Lord, 1987; 10. Harnish et ul., 1983; 11. OECD, 1981; 12. Haky and Leja, 1986; 13. DeFoe et al. 1990; 14. Gledhill et ,I. 1980; 15. DeBruijn and Hermens, 1989; 16. Brooke et al. 1990; 17. Klein et al. 1988; 18. Wolfe et al. 1980a; 19. R u u e l l and McDuffie, 1986; 20. Tanii and Hashimoto, 1982; 21. Schwarz and Miller, 1980; 22. Giam et al. 1986; 23. Hollifield, 1979; 24. C M A , 1983; 25. Boese, 1984; 26. Sullivan et al. 1981; 27. Stephen.son and Malanowski, 1987; 28. Grayson and Fosbraey, 1982; 29. Wemer, 1952; 30. G-uckel et al. 1982; 31. Veith et al. 19791); 32. Dobbs et ai. 1984; 33. Dobbs and Cull, 1982; 34. Fdssel, 1956; 35. Sears and Darby, 1982; 36. Quackenbos, 1954; 37, Ellington and Floyd, in press; 38. Cowen and Baynes, 19g0; 39. Exxon Biomedical Sciences, Inc., 1996a.
674 Other evidence indicates many of the measured water solubilities for high molecular weight phthalate esters reported in the literature are erroneously too high. First, octanol-water partition coefficients obtained by the most reliable experimental techniques (i.e. generator column or "slow-stir" shake flask methods) indicate much higher values when compared to conventional methods. Numerous studies using reliable methods (as discussed in the next section) indicate that the log Ko,vof DEHP is consistently in the 7.0-7.8 range (De Bruijn and Hermens, 1989; Brooke et al., 1990; Ellington and Floyd, in press). However, early Kowmeasurements obtained using less reliable conventional shake flask procedures yielded values two orders of magnitude lower (OECD, 1981; Brooke et al., 1990). Given the well established inverse relationship between true water solubility and Kow, these results suggest that historical water solubility determinations obtained by conventional methods are likely subject to large systematic errors.
Lack of aquatic toxicity provides additional evidence supporting the view that the true water solubilities of higher molecular weight phthalate esters are lower than most data suggest (Adams et al., 1995, Rhodes et al., 1995; Staples et al., 1996). The non-toxic nature of higher molecular weight phthalate esters implies that the true water solubility must be sufficiently low to prevent test organisms from achieving a critical body burden that elicits adverse effects. Given the imprecision of experimental measurements, coupled with the inconsistencies between experimental data and 1) theoretical predictions based on structure-activity; 2) experimental Kow measurements obtained using the most reliable methods available; and 3) aquatic toxicity information, water solubility measurements for higher molecular weight phthalate esters (Table 3) should be viewed as suspect. Predictions based on structure-activity relationships and from slow-stir measurements of log Koware believed to provide more reliable estimates than most data obtained using conventional shake flask or HPLC methods. Further research to confirm the validity of these predictions is needed.
Octanol/Water Partitionin~ The equilibrium distribution of an organic chemical between water and octanol (K~) is an important physical constant for predicting the tendency of a chemical to partition to water, animal lipids, sediment, and soil organic matter. Numerous correlations describing relationships between K~wand soil sorption, water solubility, bioconcentration, and toxicity have been developed (Lyman, 1982; DeWolfe et al., 1992; Veith et al., 1980; Gossett et al., 1983; and Geyer et al., 1984).
An early method of determining K,w was to agitate test chemical in a two-phase mixture of octanol and water and measure the equilibrium concentration of the chemical in both phases (OECD, 1981). This simple design contains a number of technical difficulties. Equilibrium must be reached and adequate separation of phases attained. The equilibrium eoneenlration of the test chemical in both phases must be below its maximum solubility in each phase. For highly hydrophobic and low solubility chemicals such as high molecular weight phthalate esters, the concentration will be substantially higher in one phase making accurate measurements in both phases dittleult. The use of radionetive test
675 chemical avoids some of these practical difficulties but introduces new problems in that the radioactivity measured at low activities may be difficult to interpret due to traces of water soluble impurities. Chemical transformations may also confound results.
A number of predictive methods have been developed to estimate octanol-water partition coefficients (Veith et al., 1979b; Lyman et al., 1982; DeKoch and Lord, 1987; McDuffie, 1981; Renberg, et al., 1985; Klein et al., 1988; Doucette and Andren, 1988; and Long, 1995). Many of these methods relate K~,,to other physical properties. Another approach is to estimate K~v based on HPLC retention time. Howard et al. 0985) reported results for a number of lower molecular weight phthalate esters using the HPLC method; however, the HPLC method could not be used to determine log K~ for the higher molecular weight phthalate esters DIOP, DIDP, DUP, and DTDP because these phthalate esters did not elute from the column under test conditions. Thus, higher molecular weight phthalate esters are too hydrophobic to quantify. Ko,, by this method.
The simplest and currently most reliable method for determining K~s for low solubility, hydrophobic compounds is the "slow-stir" method (DeBruijn and Hermens, 1989; Brooke et al., 1990; and Ellington and Floyd, in press). This procedure allows equilibrium while minimizing the formation of emulsions that can confound exTer:rnen~l results.
A summary of published Kowmeasurements ofphthalate esters are presented in Table 3. Fairly cons~ste'nt results are seen for the lower molecular weight phthalate esters (DMP, DEP, DnBP, and BBP). With increasing alkyl chain length, the log K~ increases indicating greater hydrophobicity, as expected. The median log K~ values for DMP, DEP, DnBP, and BBP were 1.61,2.38, 4.45, and 4.59, respectively. DEHP is the most widely studied phthalate ester of the higher molecular weight phthalate esters. Howard et al. (1985) estimated a log K ~ for DEHP of 7.94 using HPLC. Measurements using the slow-stir technique ranged from 7.14 to 7.45 (Brooke et al., 1990; DeBruijn and Hermens, 1989; Ellington and Floyd, in press). The U.S. EPA estimated a log K ~ of 7.54 for DEI-IP based on molecular structure using the SPARC model (Long, 1995). Log I~,,, values for higher molecular weight phthalate esters DrNP through DTDP were estimated using SPARC to be even higher, ranging from 9 to 13 (Ellington and Floyd, in press; Long, 1995). Similar estimates were obtained using the EPIWIN Model (Meylan and Howard, 1995).
Vapor Pressure
Vapor pressure plays an important role in the fate of ~gifive emissions and other releases of ph~flate esters to the atmosphere. Vapor pressure is typically deW.tined by direct pressure m~surement at elevated temperatures. Such data may be extrapolated to estimate vapor pressure at ambient temperatures using either the Clausius-Clapeyron or Antoine equations. The direct pressure measurement technique is limited by the sensitivity of the pressure measuring
676 device for poorly volatile substances. Since it is non-selective, the vapor pressure of the most volatile component of mixtures is measured.
Another method for vapor pressure measurement of poorly volatile substances is the gas saturation technique (U.S. EPA, 1980). It relies on a saturator column to generate a continuous stream of vapor saturated air at ambient temperature. The vapor is trapped and accumulated on a sorbent which is ultimately desorbed and analyzed chromatographically. This assures that the method is substance specific and not invalidated by the presence of volatile contaminants. Howard et al. (1985) utilized the gas saturation method to measure the vapor pressure of a series of phthalate esters. However, since test conditions were at 25°C, rather than elevated temperatures, the sensitivity of the method may be limited by the low air concentrations of poorly volatile chemicals.
Vapor pressure measurements reported in the literature for phthalate esters are summarized in Table 3. Measured values obtained in different studies often vary by more than one order of magnitude. However, a general trend is apparent showing that the vapor pressure of phthalate esters decline more than four orders of magnitude with increasing alkyl chain length. Comparison of measured data with the model predictions obtained by structure activity relationships included in the ASTER (Russom et al., 1991) and EPIWlN (Meylan and Howard, 1995) models often indicate more than order of magnitude discrepancies (Table 3).
ENVIRONMENTAL PARTITIONING Air-Water Partitioninc, A chemical's equilibrium distribution between water and air serves as a guide to estimate the tendency of a substance to escape from water into air. The ratio of the vapor pressure to the molar water solubility estimates the Henry's Law constant, which is the measure of the equilibrium distribution coefficient (Thomas, 1982). Henry's Law constants (H) were calculated from the recommended vapor pressure and water solubility data shown in Table 3.
For lower molecular weight phthalate esters (DIvIF, DEP, DAF, DPP, DnBP, DIBP, and BBP), H values ranged from 1.2E=7 to 8.8E-7 atm-m3/mole. Compounds with H values in the range of 1.0E=? a~n=m3/mole are generally considered to have negligible volatility (Howard et al., 1985). For all higher molecular weight phthalate esters except BOP, calculated H values ranged from about 1.7E=5 to 5.5E-4 atm-m3/mole. The higher H values are due to the greater decrease in water solubility relative to vapor pressure with increasing alkyl chain length. The H values for some of the higher molecular weight phthalate esters (DINF, DIDP, 71 IP, DUP, and DTDP) are difficult to calculate due to the extremely low vapor pressures and water solubilities that are not accurately known.
677 Vopor-Aerosol Partitioninv The distribution of organic chemicals in air between gaseous and particulate phases can be estimated by the Junge model (Eisenreich et al., 1981): 0
-
cO Po*cO
m
(I)
where 00 (phi) is the fraction of chemical on the particulate (i.e. aerosol) phase, c is a constant (approximately 0.13 for many organic chemicals), 0 is the particle surface area per unit volume and Po is the liquid vapor pressure in mm Hg at the temperature of interest. Assuming a typical suspended particulate concentration in air of 20-40 ug/m3 and a surface area of 1-3 m2/g, 0 is approximately 3 x 10"6cm-'/cm3 (Strachan and Eisenreich, 1988). By substituting this value and the vapor pressures of the individual phthalate esters reported in Table 3 into equation (1), ~ values for each phthaiate ester can be determined. Results are expressed as a dimensionless decimal fraction between 0 and I. Therefore, as 00 decreases, more chemical resides in the vapor phase. Results indicate that phthalate esters with alkyl chain lengths of less than six carbons (DMP through BBP) exist primarily in the vapor state whereas compounds with longer alkyl chains are mainly associated with aerosols (Table 4). This simple calculation does not take into account a variety of factors that can affect partitioning such as variability in particulate concentrations or surface area or the influence of temperature on the vapor pressure. Nevertheless, this calculation illustrates trends in the relative importance of atmospheric partitioning for different phthalate esters.
Partitioningbehavior influences the removal of contaminants from airvia washout by rain or snow. The washout ratio (W) is defined as the dimensionless ratioof chemical concentrations in precipitationto that in air.Washout can occur by both vapor and particle scavenging mechanisms and can be expressed as (Biddleman, 1988):
w = (1-00) ~ + 00 w
RT = (1-oo) - - f f , 00
(2)
where W, is the vapor washout ratio, RT/H in the reciprocal of the dimensionless Henry's constant at the temperature of interest, W, is the particle scavenging coefficient and 00 was defined in equation (1). Based on measurements for selected phthalate esters reported by Ligocki et al. (1985), a representative value for Wpis 20,000. Using this estimate and Hand 00 values reported in Table 4, equation (2) was used to calculate washout ratios (W) for each phthalate ester (Table 4). Limited field data indicate washout ratios for DEP, DnBP and DEHP range from 1000 to 100,000 (Arias and Glare, 1981; Ligocki et al., 1985; Atl~ and Glare, 1988; Thuren and Larsson, 1990). Thus field measurements are of the same order of magnitude as simple model predictions obtained using equation (2). One key factor contributing to the variability of field-determined washout ratios appears to be the temperature dependence of the vapor pressure. During the wanner summer months, the vapor pressure increases, which in accordance to equation (1), results in
678 Table 4. Selected Parameters Determining the Environmental Distribution of Phthalate Esters. Compound Henry's Constant (atm-m3/mol)
Ko, (L/kg) (soil/sediment)
K,~ (L/kg) suspended solids
dp
Washout Ratio (W)
DMP
1.22E-07
55 (1) 8O-36O (2)
<50,000 (8)"
0.00019
2.0E+5
DEP
2.66E-07
69 (3) 704-1726 (4)
79,400 (8) a
0.00039
9.0E+4
DAP
2.85E-07
0.0024
8.6E+4
DPP
3.05E-07
0.0039
8.0E+4
DnBP
8.83E-07
0.014
2.8E+4
1020 (10) a
0.038
1.3E+5
100,000 (8)"
0.072
3.1E+4
0.072
2.0E+3
0.80
1.6E+4
0.80
2.8E+4
0.80
1.6E+4
DIBP
1.83E-07
BBP
7.61E-07
1375 (3) 14,900 (5)
17,000 (3)
158,500 (8) ~ 39,800 (9) 1230 (10) ~ 7200 (11)
9ooo (6) DHP
4.40E-05
DnOP
1.03E-04
BOP
4.00E-07
DEHP
1.71E-05
52,600 (7) 2.0E+6 (8)"
87,420 (3) 482,000 (7) 510,000 (5) •
DIDP
286,000 (7)
DTDP
1.2E+6 (7)
1.0E+6 (8) a 398,000 (9) 22,000 (1 O)a 583,000 (11)
Calculated from Kd assuming a 0.10 organic carbon fraction. 1. Osipoffet al. 1981; 2. Banerjee et al. 1985.3. Russell and McDuffie, 1986; 4. von Oepen et al. 1991; 5. Sullivan et al. 1982; 6. Gledhi!l et al. 1980; 7. Williams et al. 1995; 8, Furtmarm, 1993; 9. Germain and Langlois, 1988; 10, Preston and A1-Omran, 1989; 11. Ritsema et al., 1989.
reduced partitioning to aerosols and hence lower washout ratios. In contrast, during the colder winter months, the vapor pressure is reduced so that partitioning to aerosols is enhanced resulting in more efficient particle scavenging and thus higher washout ratios. This temperature-dependent phenomenon has been observed in Sweden for the atmospheric removal of both DnBP and DEHP (Thuren and Larsson, 1990).
Water-Solid Partitionin~ The sorption of phthalate esters to soil, sediment, or suspended solids is partially governed by the relative hydrophobicity of the chemical. Such hydrophobic chemicals adsorb principally to the organic matter associated with
679 the solid. Adsorption is generally measured by shake flask techniques in which a solid, water and test chemical are agitated in a closed system. Complexities arise in interpreting solid sorption test results. Sorption to soil, sediment, or suspended solids is not always linear with the concentration of chemical or soil and may vary considerably with the particular solid used. A number of authors (Carlberg and Martinsen, 1982; Matsuda and Sehnitzer, 1971; Ogner and Sehnitzer, 1970) have reported that soluble soil humic material associates strongly with phthalate esters, increasing their apparent water solubility and decreasing their apparent degree of soil sorption. Russell et al. (1985) reported that biodegradation in unsterilized soil-water systems affected the results of phthalate ester soil sorption studies.
A number of authors have published soil or sediment and water partition coefficients (Table 4). Banerjee et al. (1985) reported organic carbon-normalized partition coefficients (K~s) of DMP measured at different depths in a soil core ranging from 80 to 360. Russell and McDuffie (1986) determined soil partition coefficients for a number of phthalate esters. Measured K~s ranged from 69 for DEP to 87,000 for DEHP. Glcdhill et al. (1980) showed a BBP I ~ of 9,000 while Russell and McDuffie (1986) showed a comparable K~ for BBP of 17,000 (Table 4). Sullivan et al. (1982) showed a wide range ofDEHP K~s, from 12,000 to 1 million. More recently, Williams et al. (1995) presented sediment sorption coefficients for a number of sediments and various phthalate esters (DHP, DEHP, DIDP, and DTDP). Three sediments with organic carbons ranging from 0.15% to 1.88% organic carbon were investigated. So.rption results followed the Freundlich model. Average organic carbon normalized partition coefficients (Koc) ranged from 52,600 for DHP, 482,000 for DEI-IP, 286,000 for DIDP, and 1.2E+6 for DTDP. Williams et al. (1995) also noted an inverse dependence of experimental K~s with increasing solids concentration. The authors calculated particle-corrected I ~ values ranging from one to three orders of magnitude higher than the uncorrected experimental I ~ values. It was concluded that failure to consider particle effects with compounds such as phthalate esters, would result in significantly overestimating truly dissolved sediment porewater concentrations.
Several authors have examined the dissolved versus suspended particulate-bound fraction ofphthaiate esters in surface water samples. Germain and Langlois (1988) collected surface water samples from the St. Lawrence River. Both dissolved phthalate ester and phthalate ester bound to suspended particulate matter-bound (SPM) were analyzed. The SPM concantration was estimated to be 3.0 mg/L. About 14% of the total DnBP concentration was sorbed to SPM and 86% was dissolved. For DEHP, 53% was SPM-bound while only 47% was dissolved. No other phthalate ester had detectable SPM-bound fractions. Ritsema et al. (1989) separated SPM by centrifugation from surface water samples collected from Lake Yssel and the Rhine River (Netherlands). SPM ranged from 4-100 mg/L. Based on the geometric mean of th© range of SPM values, 98% of the DBP present was dissolved while only 2% was SPM-bound. For DEHP, 33% was estimated to be dissolved while 67% was estimated to be SPM-bound. Preston and AI-Omran (1986, 1989) collected a series of surface water samples from the River Mersey Estuary in the UK. SPM concentration reportedly was a relatively high 1524 mg/L. For DnBP, the authors reported 66% to 86% of the total was dissolved, while 14%
680 to 34% was particulate bound. Weekly samples have been collected for several years from the source and mouth of the Niagara River connecting two Great Lakes between the US and Canada (Data Interpretation Group, 1986-1993). SPM-bound and dissolved concentrations for several phthalate esters were measured in this monitoring program. The reports indicate that SPM averages 7.83 mg/L in the river samples. Dissolved DnBP averages 83.5% while SPM-bound DnBP was 16,5%. For DEHP, 73% was dissolved and 27% was particulate bound. Furtmarm (1993) reported similar findings. For the low molecular weight phthalate esters (DMP to Bttl'), 15-17% was particulate bound while for DEHP and Dnt)P 53-74% was particulate bound. The amounts of dissolved DEHP determined in these studies are consistently less than predicted from simple two-phase partitioning models. The differences are attributed to, in part, separation techniques that do not distinguish chemical complexed to organic colloids from truly dissolved phase chemical. A summary of partition coefficients between SPM and natural waters obtained from these studies are presented in Table 4.
It is generally accepted that only the truly dissolved phase of a non-polar organic chemical is bioavailable (Steen et al., 1980; McCarthy, 1983; Rodgers et al., 1985; Staples et al., 1985; McCarthy and Black, 1988; DiToro et al., 1991). Therefore, the partitioning of phthalate esters to colloidal and particulate organic carbon should be considered in analysis of field samples. However, most historical measurements are based on total concentrations that fail to differentiate free and complexed forms. Consequently, available field data may significantly overestimate bioavailability especially for the more hydrophobic phthalates that are expected to exist principally in the environment as complexed forms.
ENVIRONMENTAL TRANSFORMATIONS
Hvdrolvsis Phthalate esters are susceptible to hydrolysis, however at slow rates. The products of hydrolysis are an acid and an alcohol. Phthalate esters can undergo two hydrolytic steps, producing first the mono-ester and one free alcohol moiety and a second hydrolytic step creating phthalic acid and a second alcohol. Ester hydrolysis may be either acid or base catalyzed, with, in some instances, metal ions, anions, or organic materials serving as catalysts (Harris,
1982a). Phthalate esters are hydrolyzed at negligible rates at neutral pH (Table 5). Acid hydrolysis of phthalate esters is possible, but is estimated at four orders of magnitude slower than alkaline hydrolysis rate constants (Mabey et al., 1982). As an example of the limited extent of acid hydrolysis, Thuren and Sodergren (I 987) used extraction into
681
Table 5. Summary of Abiotic Degradation Half-lives of Phthalate Esters.
• b
Phthalate Ester
Aqueous Hydrolysis Half=Lives (years)'
Atmospheric Photooxidation Half=Lives (days) b
DMP
3.2
9.3-93
DEP
8.8
1.8-18
DAP
--
0.04-0.4
DPP
--
0.9-9.0
DnBP
22
0,6-6,0
DIBP
--
0.6-6.0
BBP
>0.3
0.5-5.0
DHP
--
O.4-4.0
DnOP
107
0.3-3.0
BOP
--
0.4-4.0
610F
--
0.2-4.0
DEHP
2000
0.2-2.0
DIOP
157
0.3-3.0
DINP
--
0.2-2.0
DIOP
--
0.2-4.0
711P
--
0.2-4.0
DUP
--
0.2-2.0
DTDP
--
0.2=2.0
Hydrolysis rate constant to monoester at pit--7, 25 C (Wolfe et aL 1980a), except for BBP (Gledhill et al., 1980) and DnOP (Howard, 1991). Predicted half-lives obtained by the Atmospheric Oxidation Program (Atldnson, 1988).
hydrated sulfuric acid to extract phthalate esters in a clean-up step for analysis and obtained 100% recoveries. Wolfe et al. (1980a) measured alkaline hydrolysis rate constants for a number of phthalate esters from free energy relationships. Hydrolysis rates decrease and corresponding half-lives increase with increasing alcohol chain length. Half-lives ranged from about 3 years for DMP to 2000 years for DEHP. Gledhill et al. (1980) estimated a hydrolysis half-life of>100 days for BBP. Hydrolysis is unlikely to be an important fate process for phthalate esters under typical environmental conditions (Schwartzenbach et al., 1992).
682 Photodegradation Aqueous photolysis occurs through absorption of UV light from sunlight in the region of 290-400 run. Shorter wavelengths are attenuated by passage through the atmosphere and water column. Longer wavelengths lack sufficient energy to break covalent bonds (Harris, 1982b). Photolysis can be mediated via either direct or indirect mechanisms. The mechanism of photolysis may be either through direct absorption of UV radiation by the chemical or by absorption of UV radiation by natural substances such as water with the formation of activated species such as singlet oxygen or hydroxy radicals that react with phthalate esters.
Few studies on phthalate ester photolysis are available. Gledhill et al. (1980) reported that a 1 mg/L solution of BBP exposed to sunlight for 28 days resulted in less than 5% degradation. It was concluded that BBP has an aqueous photolysis half-life of>100 days. Wolfe et al. (1980b) estimated a maximum near-surface phthaiate ester photolysis half-life of 144 days for phthaiate esters. This estimate was based on unpublished experiments with DMP spiked into surface water that was irradiated with sunlight for one week. Photooxidation ofphthalate esters in sunlit surface waters does not appear to represent an important transformation process. Based on the data of Wolfe et al. (1980b), Howard (1991) estimated that aqueous photooxidation half-lives range from 2.4 to 12 years for DEP and DnBP and from 0.12 to 1.5 years for DEHP.
In contrast to the minor role of photodegradation in natural waters, these reactions appear to be much more important in the atmospheric fate of phthaiate esters. Reaction with hydroxyl radicals is generally the most important photodegradation process for organic chemical pollutants in the atmosphere (Kelly et al., 1994; Atkinson, 1988; Meylan and Howard, 1993). Predicted photooxidation half-lives for phthalate esters based on hydroxyl radical attack were obtained from structure activity relationships contained in the Atmospheric Oxidation Program (AOP) (Meylan and Howard, 1993). Reported half-lives are specified as a range to indicate differences that are expected due to varying hydroxyl radical concentrations between pristine (3x10 "~radicals cm3) and polluted (3xI0 6 radicals cm-3) air. Results indicate that as alkyl chain length increases, susceptibility of phthalate esters to photooxidation also increases (Table 5). While limited experimental data are available for comparison with AOP predictions, one study reported an experimental atmospheric half-life for DEHP of about 1 day (Behnke et al., 1987), which agrees with model predictions.
a~air,~.gzl~a~ Biodegradation is a critical process affecting the environmental fate of phthalate esters. Considerable research has been conducted on the biodegradability of this chemical class over the last few decades. Microbes from diverse habitats have been shown to degrade phh'~alat¢esters and resulting intermediates (Stahl and Pessen, 1953; Mathur and Rouatt, 1975; Englehardt et al., 1975; Keyser ¢t al., 1976; Nagata et al., 1976; Klausmeicr and Osmon, [976; Kuran¢ ¢t al., 1977a, b; Englehardt et at., 1977; Englehardt et al., 1978; Nakazawa and Hayashi, 1978; Shimada and Shimahara, 1978; Yoshida
683 et al., 1979; Taylor et al., 1981; Benckiser and Ottow, 1982; Eaton and Ribbons,' 1982; Williams and Dale, 1983; Karegoudar and Pujar, 1984; Kurane, 1986; Gibbons and Alexander, 1989; Nomura et al., 1989; Yan et al., 1995). Representatives from both aerobic and some anaerobic environments include gram positive and gram negative bacteria and actinomycetes. Although some individual microbes are capable of completely mineralizing phthalate esters, more efficient metabolism appears to result from mixed microbial populations, typically found in the environment (Engelhardt et al., 1975; Aftring et al., 1981; Kurane, 1986).
Research suggests that the metabolic pathway for the microbial metabolism of phthalates under both aerobic and anaerobic conditions begins by ester hydrolysis to form monoester and corresponding alcohol. Under aerobic conditions, further enzymatic degradation of the monoester proceeds via phthalie acid by either a 3,5 or 4,5 dihydroxyphthalate pathway to procatechuate. Aromatic ring cleavage of procatechnate can then occur via either an ortho pathway that results in the formation of pyruvate and oxaloacetate or a meta pathway yielding a 13-ketoadipate that is further degraded to acetyl CoA and succinate (Eaton and Ribbons, 1982; Nomura et al., 1989; Kohli et al., 1989). Although less is known about the pathways of anaerobic catabolism, it appears that the monoester is degraded to phthalic acid and then further degraded by the same pathway used for benzoate (Afh-ing et al., 1981, Ejlertsson and Svensson, 1995). Benzoate has been shown to be readily degraded anaerobically (Battersby and Wilson, 1988; 1989; Shelton and Tiedje, 1984). Based on the recent comprehensive review by Ejlertsson and Svensson (1995), the proposed biodegradation pathway for phthalate esters is presented in Figure 1.
Numerous studies have been conducted to evaluate the primary and ultimate biodegradability of phthalates using both standardized and non-standardized test protocols. For the purpose of this review, studies can generally be divided into three types based on the nature of the test matrix: inoculum tests, freshwater and marine water tests, and soil/sediment tests. Inoculum tests involve spiking the test chemical to an aqueous solution containing inorganic nutrients and a microbial inoculum (e.g. wastewater effluent, sludge) obtained froma domestic or bench scale wastewater treatment system. The microbial inoculum used in such tests may or may not be acclimated to the test chemical. Acclimated inocula typically reduces or eliminates the time lag observed before the onset of biodegradation. Freshwater/marine and soil/sediment tests typically involve following the degradation of test chemical added to flasks containing natural water or soil/sediment. Alternatively, microcosm studies have also been used to assess the potential biodegradability of phthalate esters in natural waters and sediments. Microcosm studies have been performed with DEP (Lewis et al., 1985), DnBP (Tagatz et al., 1986); BBP (Adams et al., 1989; Adams and Saeger, 1993); DnOP (Sanborn et al., 1975) and DEHP (Sodergren, 1982; Perez et al., 1983; Davey et al., 1990). Field studies have also been conducted to assess phthalate ester biodegradation in soil (Falrbanks et al.~ 1985; Schmitzer et al., 1988). Such tests share the common
684
[~C 0 0 R COOR Dialkyl Phtha[ate
[~CO
0R COOH
Mono Alkyl Phthalate
Ic°? H
~:oo. or
I
OH
OH 3.4 Oihydroxy Phthalate
~'~/COOH A
=CO CoA I COOH
O
OH 4,5 Dihydroxy Phthalate
I
Phthalic Acid
ATP l + Co AADP
v'~COz OCoA
I
Benzoyl - CoA
otochatechuate
mete i ortho I cleavage
~1~0 0 H CHO /HOOC 2-hydroxy4-carboxym uconic sere leldehyde
oA r ' 1 Cycionex. 1. erie c a r b o x y l - CoA
I
~j~O0 H
U , ~ COOH COOH 13.Carboxy-CisCis-Muconate
[ . ~ (~1-~ 2-H y~roxycyclohexane LVJ Co,~arboxyl - CoA
3 Acetate + ~ 3Hz + CO2
¥
Acetate ÷ CO z + Succinate
T
.
.
.
.
CO CoA 11 CO CoA L~ Pimeiyl di.CoA
2 Pyruvate + CO z
Figure 1. General biodegradation pathway for phthalate esters in the environment.
685 objective of assessing the potential of indigenous microbial populations to degrade the test chemical. Consequently, an inoculum is not added nor is the influence of acclimation on the microbial consortia considered in these studies. Inorganic nutrients, however, are sometimes added.
An extensive summary of biodegradation test results for phthalate esters is provided in Table 6. Test conditions including the use of acclimated inocula, nutrient addition, temperature and initial phthalate concentration are provided for comparison. Tests assessing primary degradation are based on measuring the disappearance of parent phthalate by a specific analytical method (e.g. GC or HPLC analyses). In comparison, ultimate biodegradation is assessed by measuring carbon dioxide evolution or oxygen uptake under aerobic conditions or methane evolution under anaerobic conditions. For a number of studies compiled in Table 6, radiolabeled compounds were employed.
Aerobic degradation studies using inoculum tests have been performed by Saeger and Tucker (I 973 ); Barth and Bunch (1979); Urushigawa and Yonezawa (1979); Tabak et al. (1981); Patterson and Kodukula (1981); Sugattet al. (1984); O'Grady et al. (1985); Nyholm (1990); Struijs and Stolenkamp (1990); Aichinger et al. (1992); Association of Plasticizer Industry Japan (1994); Scholtz (1994a,b); and Exxon Biomedical Sciences, Inc. (1995a, b, 1996). Primary degradation for the lower molecular weight phthalates DMP, DEP, DBP and BBP occurred rapidly, typically exc~ding 90% degradation within a week even if unacclimated inocula were used. Most higher molecular weight phthalate esters demonstrated primary degradation in excess of 90% al~er 12 days using acclimated inocula. Slower rates of degradation we~ observed for unacclimated test systems. Ultimate biodegradation data for various phthalate esters are plotted in Figure 2. These results are based on 28-day experiments except for the data of Aichinger (1992). For this study, biodegradation was calculated fi'om the estimated, time-independent, growth yield obtained using a kinetic model. The studies by Saeger and Tucker (1976); Sugatt et al. (1984); Nyholm (1990); Struijs and Stolenkamp (1990); Scholtz (1994a, b); and Exxon Biomedical Sciences, Inc. (I 995b), measured CO2 evolution endpoints, whereas the studies by Aichinger et al. (1992); Association of Plasticizer Industry Japan (1994); and Exxon Biomedical Sciences, Inc. (1995a, 1996) measured oxygen uptake.
Figure 2 shows that aerobic biodcgradation test results for different phthalatc esters obtained from different studies are generally in good agreemcm. Consistent differences in results using tests with acclimated and unacclimated test inoculum are not obvious. However, results for DEHP reported by Struijs and Stolenkamp (1990) appear to be inconsistently lower than other studies. This discrepancy may be explained by the method in which DEHP was added to the test system. In this study, DEHP was impregnated into filter paper which was then introduced into the test system. As a result, the bioavailability of impregnated DEI-IP is likely limited by the desorption rate of DEHP from the filter paper. If the result from this study is excluded, all phthalate esters investigated undergo 260% ultimate
AS; WW SL
Aerobic-Ultimate
SG(4); SM SG SL SD (2) SG (2) $D(2)
Anaerobic-Primary
FW FW FW FW ,SL SL(2); SM SG(3); SL(3) FW WW FW WW SL(2); SM SL, WW
SM; AS SM; WW SM; WW SM; AS; SL
AS AS SM; AS
Aerobic-Primary
Test Matrix
N/N N/N
N/Y N/N N/N N/N N/Y N/N
Y/Y N/Y Y/Y Y/Y N/Y N/NR Y/Y N/N N/N N/N N/N N/N N/N N/Y N/N Y/Y N/N N/N N/N N/Y
Acclimation/ Nutrients
25 20
35 37 30 35 35 35
NR 25 22 22 25 NR 22 20 4 NR ? 30 25 NR 20 25 25 NR 25 20
Cone. (ppm)
100 27
30-81 0.5-10 500 81 20 50
0.1 - 1.0 20 I-3 3 5 5 20 0.0035 0.0035 0.0003 ? 500 1 703-1500 0.00074 0.00074 0.0018 NR 1 0.020
Dimethyl Phthalate
(*c)
Temp
Initial
28 3.75
7-70 7 30 56-96 70 56-96
1 7 1 1 7 7 28 3 I0 1 14 10 4 76-97 5 17 7 7 4 2
(Days)
Test Duration
Table 6: Summary of Phthalate Biodegradation Literature
90-98 50
100 77' 63 18-40 >70 18-40
100 48" >81 >90 100 >90 >99 100 0 97 ,,b 20-100 91 85 - 100 64-94 95 64 100 100 86-100 99.8
(%)
Degradation
15 16
17 18 10 13 34 13
t 2 3 3 4 5 6 7 7 8 9 10 11 12 14 14 14 14 11 71
Reference
O~ O0 O~
N/Y N/N Y/Y N/N N/N N/N N/Y
SG(3); SL(3)
FW WW FW WW SL(2); S M SL; WW
SC_~4);SM
N/Y
25 NR 22 20 4 20
N/Y N/NR Y/Y N/N N/N N/N N/N N/N
35
20 25 25 NR 25 20
NR
15-32
25 22
Diethyl Phthalate
37 37 35 35 35
22 25 20
N/Y Y/Y
Anaeroblc-Primary
Initial Cone. (ppm)
20 - 50
0.00086 0.00086 0.0024 NR I 0.020
! 000
20 I-3 3 5 5 20 0.0025 0.0025 0.19 25 0.02-2.2
20 20 - 200 81 81 50
20 55 0.020
Dimethyl Pbthalatc (cont.)
Temp (°C)
AS SM; AS SM; AS SM; WW SM; WW SM; AS; SL FW FW FW FW SL
Aerobic-Primary
SG; SM SG; SM SG(2); SM SG; SM SG(3)
N/Y N/Y N/NR N/Y N/Y
Y/Y Y/Y N/Y
SM; AS; SL SM;MI SL; WW
Anaerobic-Ultimate
Acclimation/ Nutrients
Test Matrix
7 - 70
5 17 7 7 41 2
287
7 1 ! 7 7 28 3 I0 ! 14 5 I- 129
70 140 7 77 32 - 56
28 NA 4
Test Duration (Days)
Table 6: Summary ofPhthalate Biodegradation Literature
100
95 0 99 98 86-100 100
95
36" >94.8 >90 ! 00 > 90 >99 100 0 76- l 0(P 14-100 19-99
82 94 - 100 58 - 88 41 58 - 88
86 96 c 99.6
(%)
Degradation
17
14 14 14 14 II 71
!2
2 3 3 4 5 6 7 7 24 9 25
34 21 22 23 17
6 19 71
Reference
O, oo
Aerobic-Primary AS SM ; AS SM; AS SM; AS; SL SM; WW SM; WW FW FW FW FW FW FW FW MW FW FW; SD FW
SG(3) SC
Anaerobic-Ultimate SG; SM SG(2); SM
SL; WW
Y/Y Y/Y Y/Y N/Y N/NR N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N
N/Y
N/Y N/NR N/Y N/Y
Y/Y Y/Y N/Y
N/N N/Y N/Y
SG SG(2) SC
Aerobic- Ultimate SM; AS; SL SM;MI
Acclimation/ Nutrients
Test Matrix
20-200 77 50 50-250
20 37 0.020
0.5 - 10 20 50-250
20 1-3 3 20 5 5 1 0.05 0.00325 0.00325 0.001 25
< 1.6 < 3.8 0.4-0.55
25 22 22 22 25 NR 25 NR 20 4 NR
28 28 25
Di-n-Butyl Phthalate
37 35 35 37
22 25 20
37 35 37
Diethyl Phthalate (cont.)
Temp (*C)
Initial Cone. (ppm)
7 ! 1 28 7 7 4 3 7 10 I 14 I0 10 7 7 9-11.5
17 28-56 32-70 100
28 NA 6
7 70 60- 100
Test Duration (Days)
Table 6: Smmnm~ o f Phthalate Biodegradation Literature
60 > 95 a 90 90 i00 a90 99 > 90 100 0 82" 33-100 45-95 55 49 32-49 80-100
33-56 0-32 0-76 5-90
95 93 ¢ 99.2
64' >90 85-100
(%)
Degradation
2 3 3 6 4 5 26 27 7 7 8 9 28 28 29 29 30
21 22 17 75
6 19 71
I8 34 75
Reference
O~ OO Oo
FW; SD MW M W ; SD MW SL(2); S M SD SD SL SL(2) SL(2) SL(2) SL SG(3); SL(3) FW WW SL; WW SL; GW
Aerobic-Ul~mme SM; AS; SL
SL SC SL; GW
SDO); SM
SG; SM SO (2) SO (3) SO; SM SO SG;SM
Y/Y
N/Y N/Y N/Y N/Y N/N N/Y N/Y N/N N/N N/N
N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N N/Y N/N N/N N/Y N/N
Matrix
Anaerobic-Primary
Acclimation/ Nutrients
Initial Conc. (ppm)
20 10
22
20
20 20 20 50 0.5-10 56 56 500 500 5
NR 0.024 5
NR
35 35 35 35 37 30 30 30 ! 8-68 10
0.4-0.55 0.3-0.5 0.3-0.5 0.3 I ! 0- 1000 10- 1000 500 200 800 1600 0.02-2.2 265-596 0.0016
25 25 25 25 25 28.5 28.5 30 21 2! 21 15-32 NR 25
Di-n-Butyl Phthalate (cont)
Temp (°C)
28
70 28 17-70 68 7 63 22-42 30 30 35-166
2 15
7
7-9 7-17 2-13 1 3 42-56 42-56 !0 182 182 182 34-129 76 7
Test Duration (Days)
Table 6: Summary of Phthalate Biodegradation Literature
57
100 100 80-100 100 83" 81 > 95-100 66 79 50~
99.3 50~
75
!00 95-100 I00 ! O0 97-100 59-96 52-89 92d 66-98 35-43 40 88-97 51-79 59
(%)
Degradation
6
34 34 17 i7 18 35 35 10 36 48
7I 48
14
33 33 25 12 14
33
30 30 30 3I II 32 32 I0
Reference
,4D
Aerobic-Primary SM; AS; SL SM; AS SM; WW SM; WW FW FW FW FW FW; SD FW; SD FW
SL; SM SL; WW
sG(12); SM
SG; SM SG; SM SO(2); SM SG(2); SM
A naerobic- Ultimate
FW; SD FW; SD FW; SD FW; SD SL; WW
SL(I); WW
N/N N/N N/N
Y/Y N/Y N/Y N/NR N/N N/N N/N
N/Y N/Y N/NR N/Y N/Y N/N N/N
N/Y Y/Y N/N N/N N/N N/N N/N N/N N/N N/N N/Y
SM; AS
SM;MI SL SL(4) SL(I)
Acclimation/ Nutrients
Test Matrix
20 20-100 72 72 50 1 1000
22 8 I 1000 1000 1000 0.082-8.2 NR NR NR 0.024
22 NR 25 NR NR NR 20 4 NR 20 20
28 I 7 7 9 7 7 10 7 2
5
0.00162
70 140 14-21 60 56 30 80
28 NA 30 80 80 53 14 14 14 14 12
Test Duration (Days)
20 NR 5 5 1 I 0.0035 0.0035 1 0.01-0.1
Butylbcnzyl Phthalate
35 37 35 35 35 22 23
21 25 22 23 30 4 22 12 28 5 20
Di-n-Butyl Phthalatc (cont.)
Temp (°C)
Initial Cone. (ppm)
Table 6: Summary of Phthalate Biodegradation Literature
100
87 96 100 a 90 99 100 100 0 95 47-60*
80 100 32-85 24-59 3 2 - 85 98 84
81 92 c 97 84-99 83 2 71-85 80 95 50 99.8
(%)
Degradation
14
6 42 4 5 42 43 7 7 44 45
34 21 22 41 17 38 39
37 19 38 39 39 39 40 40 40 40 71
Reference
'*,D 0
SM; AS
Aerobic-Primary
SG(2); S M FW; SD
so(2); S M
SO; SM SM;SG?
Anaerobic-Ultimate
SM; AS; SL SM; AS SM;MI FW; SD FW; SD FW; SD SM; AS AS; WW
Aerobic-Ultimate
SG; SM SG; SM SG; SM SG SG SG(3) SD(3); SM FW; SD
Y/Y
N/Y NR/NR N/Y N/Y N/N
Y/N Y/Y Y/Y N/N N/N N/N N/N N/N
N/Y N/Y N/Y N/N N/N N/Y N/Y N/N
Y/Y N/N N/N
WW FW WW
Anaerobic-Primary
Acclimation/ Nutrients
Test Matrix
Initial Cone. (ppm)
22
Dihexyl Phthalate
35 NR 35 35 NR
22 NR 25 NR NR 20 N/R 25
35 30 35 37 35 35 30 NR
25 25 NR
1-3
20 NR 68 50 I
20 20 9-45 l 1 0.Ol-O.I N/R 100
20 62 50 4 20 20 62 I
0.00162 0.0067 NR
Butylbenzyl Phthalate (cont.)
Temp (°C)
1
70 28 28-56 32-56 28
28 27 NA 28 28 30 28 14
48 63 48 7 28 42-70 42- 100 2
i7 7 7
Test Duration (Days)
Table 6: Summary ofPhthalate Biodegradation Literature
> 93
I00 <10 0-24 0-100 45
43 95 91-93 ~ 50 51-65 10.4 96 81
100 52 100 66' 90 75-98 100 50'
95 100 I00
(%)
Degmdmion
3
34 44 22 17 46
6 42 19 46 44 45 44 15
34 35 17 18 34 17 35 46
14 14 14
Reference
O~
SG; SM
Anaerobic-Ultimate
SL; WW
SM;MI
Aerobic-Ultimate
SG; S M SG SO(3) SG
Anaerobic-Primary
FW FW FW SL SC FW WW
SM; WW SM; WW
AS
Aerobic-Primary
SM; AS; SL SM; AS
N/NR
Y/Y N/Y
N/Y N/N N/N N/N
N/Y N/Y N/NR YINR N/N N/N N/Y N/N N/N N/N N/N
Y/Y N/Y
Y/Y N/Y N/N
SM; AS (cont.) SM; AS; SL AS FW
Aerobic-Ultimate
Acclimation/ Nutrients
Test Matrix
35
25 20
35 37 35 35
25 25 NR NR 20 4 26 15-32 18-69 25 NR
20 35
3 20 20 25
68
8-39 0.020
50 4 20 20
20 5 5 5 0.0025 0.0025 2.5 0.02-2.2 500 0.0034 NR
Di-n-Oetyl Phthalate
22 22
22 25 ?
DihexylPhthalate(cont.)
Temp (°C)
Initial Cone. (ppm)
56
NA 20
70 32 70 70
7 7 7 7 10 I0 5 34-96 30 7 7
28 28
I 28 7 14
Test Duration (Days)
Table 6: Summary of Phthalate Biodegradation Literature
0
85-92 c 52.3
5 0 12-44 30
10~ 0 0 92 85 0 50 "d 80 31 100 97
77 80
a 90 > 99 38" 30-95
(%)
Degradation
22
19 71
17 18 17 34
4 5 5 7 7 47 25 36 14 14
6 70
3 6 2 9
Reference
tO
SM; AS SM; AS SM; AS SM; AS SM; AS SM; AS; SL SM; WW SM; WW SM; WW SM; WW FW
Aerobic-Primary
SM; AS; SL
Aerobic-Ultimate
SM; AS SM; AS SM; AS; SL
Aerobic-Primary
SM; AS; SL
Aerobic-Ultimate
SM; AS SM; AS SM; AS; SL
Y/Y Y/Y Y/Y Y/Y Y/Y Y/Y N/Y Y/Y N/NR Y/NR N/N
Y/Y
Y/Y Y/Y Y/Y
Y/Y
Y/Y Y/Y Y/Y
Initial Conc. (ppm)
50 20
22
22 22 22
20
1-3 3 20
Butyl2-Ethylhexyl Phthalate
35 35
Di-n-OetylPhthalate(cont.)
Temp (°C)
I-3 3 20
22 15 NR 22 22 22 25 25 NR NR NR
3.3-33 3.3-33 3.3 I-3 3 20 5 5 5 5 l
22 20 Di (2-Ethylhexyl) Phthalate
22 22 22
Di (n-Hexyl, n-Octyl, n-Decyl) Phthalate
N/Y N/Y
SG(14); SM SG; SM
Aerobic-Primary
Acclimation/ Nutrients
Test Matrix
I 1 1 I 3.5 28 7 7 7 7 5
28
! 3 28
28
1 3 28
56-70 70
Test Duration (Days)
Table 6: Summary of Phthalate Biodegradation Literature
31-88 22-36 74 81.5 ~ 90 > 99 0 75 0 95 40
87
60 ~ 90 > 99
90
60 ~ 90 > 99
0-13 13
(%)
Degradation
49 49 43 3 3 6 4 4 5 5 42
3 3 6
6
3 3 6
17 34
Reference
~O
SG; SM SG SG(2) SG $L SD(2); SM SD LL SL (2)
N/Y N/N N/N N/N N/N N/Y N/N N/Y N/N
N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N N/Y N/Y N/N N/Y N/Y N/Y N/Y N/Y N/N N/N N/N N/N N/N
FW FW FW FW FW FW MW MW MW SL; SM SL; SM SG(3); SL(3) SL SL; SC SL; SC SL; SC SL; SC SL; SC FW FW WW SL $L
Anaerobic-Primary
Acclimation/ Nutrients
Test Matrix
(Days)
(ppm)
30 37 35 35 30 30 NR 30 20
7.8-78 4 20 20 500 7.8-78 NR 78 1
25 1 0.05 5 20 0.0035 4 0.0035 --25 . . . . . . . . . . . . 18 0.001-0.1 1 0.001-0.1 NR 22,000 NR 22,000 NR 372-578 30 500 22 1230 22 4260 22 17700 22 21900 N/R 2000 N/R 0.05 25 0.0051 NR NR NR 5 NR 250
63 32 70 70 30 365 14 61 1O0
20 5 10 10 14 10 10 2 28 30 24-30 76-97 30 70 70 70 70 190 2 7 7 80 80
Test Duration
Initial Conc.
Di (2-Ethylhexyl) Phthalate (cont.)
Temp (°C)
Table 6: Summary o f Phthalate Biodegradation Literature
0 0 0-28 10 33 0-14 0 1 64-97
> 99 50 37 0 31-88 35-75 26 50 50 0 > 90 60-94 33-92 80-86 94 32-69 14-37 85 10 79 33 85 50
(%)
Degradation
35 I8 17 34 10 35 52 52 72
26 38 7 7 9 28 28 50 50 51 51 12 10 20 20 20 20 20 27 14 14 73 73
Reference
SL (2) SL (I) SL (I)
FW FW FW; SD SL; WW SL; SM SL;SM SL SL; SG SL; MI SL SL SL SL
SL(3)
AS; MW MW FW FW FW FW SL(3)
Y/Y Y/Y N/Y Y/Y N/Y N/Y N/Y N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N N/Y N/N N/Y N/N N/N Y/Y N/N N/N N/N N/N
N/N N/N N/N
Aerobic-Ultimate SM; MI SM; AS; SL SM; AS SM; AS SM; AS SM; AS SM; AS
Acclimation/ Nutrients
Test Matrix
Initial Cone. (ppm)
25 22 2! 20 20 22 25 25 <10 >20 29 29 29 29 N/R N/R NR 29 l0 20.5 NR NR 22 21-24 NR 22 NR NR 20-24
13.5 14.1 12.8
8-40 20 20 20 50 46 100 i00 0.002 0.002 I 0- 100 ppt 0.1 - I ppb 1- 10 ppb 10-100 ppb 0.5 10 0.1 20 ppt-200 ppb 1.4 0.1 NR NR I 2-20 l 0.029 1 2 2
! 1 1
Di (2-Ethylbexyl) Pbthalate (cont.)
Temp (°C)
28 28 27 28 28 28 14 14 14 40 40 40 40 28-63 63 32-34 28 27 28 14 14 30 146 20 33 111 446 7
NA
100 100 100
Test Duration (Days)
Table 6: Summary ofPhthalate Biodegradation Literature
38 61 12 55 59 86 50 22 93 98 2
0-80
4.5 • 63 80 29 11-16 33-69 71.2 42.3 34.9 62.3 43-58 22-34 4-28
82 86
90-95 c 86
33-89 96-99 29-72
(%)
Degradation
19 6 53 42 54 55 26 15 56 56 57 57 57 57 74 74 58 59 60 12 51 51 38 61 62 63 63 63 64
72 72 72
Reference
t~
SM; AS; SL SM; AS
Aerobic-Ultimate
SM;AS SM;AS SM; AS; SL
Aerobic-Primary
SG;SM SG;SM SG;SM SG;SM SG(3);SM SD
SG;SM
SG; SM SG;SM SG;SM SG;SM SG(2); SM
A naerobic- Ultimate
SC
SL (2)
FW; SD FW; SD SD SL (2)
Y/Y N/Y
Y/Y Y/Y Y/Y
N/Y N/Y N/Y N/Y N/? N/Y N/? N/? N/? N/? N/Y N/N
N/N N/N N/N N/N N/N N/N N/N N/N N/N N/N NN
SL SL SL FW; SD
FW; SD
Acclimation/ Nutrients
Test Matrix
Initial Cone. (ppm)
22 25
22 22 22
20 100
1-3 3 20
20 20 100 200 68 68 34 68 136 271 50 I
2.6-14 0.1 0.1 0.018 0.018 0.018-10 0.018 NR ! 1 500
Di-lsooctyl Phthalate
35 37 37 37 35 35 35 35 35 35 35 22
15-30 22 35 5 12 22 28 20 20 13.5 18-69
Di (2-Ethylhexyl) Phthalate (cont.)
(*C)
Temp
28 28
I 4 28
70 140 140 140 56 77 60 60 60 60 32-70 3O
!i5 33 56 28 28 28 28 28 100 100 30
Test Duration (Days)
Table 6: Summary of Phthalate Biodegradation Literature
57 56
84.5 a 90 > 99
9 50 7 4 0 0 48 26 19 6 0-9 0
68 23 12 I 3 6-24 10 20 2.6-47 12-33 41
(%)
Degradation
6 68
6
13 13
34 21 21 21 22 23 41 41 41 41 17 38
65 66 66 40 40 40 40 67 72 72 36
Reference
~D
SM; AS; SL SM; AS Ski; AS
Aerobic-Primary
SM; SL; AS SM; AS
Aerobic-Ultimate
SM; AS; SL SM; AS FW FW WW
Aerobic-Primary
SM; AS; SL SM; AS SM; AS SM; AS FW; SD
Aerobic-Ultimate
SM; AS SM;AS SM;AS FW
Y/Y N/Y Y/Y
y/Y N/Y
Y/Y N/Y N/N N/N N/N
Y/Y N/Y N/Y N/Y N/N
Y/Y Y/Y Y/Y N/N
N/N
FW; SD
Aerobic-Primary
Acclimation/ Nutrients
Teat Matrix
Initial Cone. (ppm)
20 45 35 100 0.02-10 NR
20 3.3-13.3 1 0.00052 NR
22 NR 22
20 3.3-13.3 !-3
22 20 22 46 Di(Heptyl, Nonyl, Undecyl)Phthalate
22 NR NR 25 NR
Di-Undecyl Phthalate
22 22 22 25 22 28
22 22 22 25
I-3 3 20 1
NR 0.002-10 NR
Di-lsononyl Phthalate
12 22 28
Di-lsooetyl Phthalate(eonL)
Temp (*C)
28 1 1
28 28
28 I 5 7 7
28 28 28 28 28 28
I 4 28 12
28 28 28
Test Duration (Days)
Table 6: Summary ofPhthalate Biodegradation Literature
> 99 51 65
76 57
> 99 29-45 20 100 76
62 70 57 71 < 1-8 2
68 ~ 90 > 99 > 95
2 < 1-10 4
(%)
Degradation
6 43 3
6 55
6 42 42 14 !4
6 55 69 26 40 40
3 3 6 26
40 40 40
Reference
"..,I
• b ° d " (#) AS MI
Y/Y
Y/Y Y/Y
Y/Y N/Y N/Y N/N
Y/Y Y/Y
Y/Y Y/Y
Y/Y N/N
Temp (*C)
3 1
22
22 22
22 22 25 25 Di-Tridecyl Phthalate
22 22
20
20 I-3
20 48 100 100
20 1-3 3
22 20 NR 20 Di-lsodeeyl Phthalate
22 NR
Di (Heptyl, Nonyl, Undecyl) Phthnlate (cont.)
Acclimation/ Nutrients
Calculated from reported first-order half lives. Primary removal process assumed to be biodegradation. Calculated from reported kinetic parameters. Includes losses due to other potential removal mechanisms (e.g., sorption). Chemical impregnated into filter paper. # of test matrices investigated e.g., SL(4) indicates 4 soils were tested. Activated Sludge Microbial innoculum from batch reactor fed with test compound as sole carbon source.
SM; AS; SL
Aerobic-Ultimate
SM; AS; SL SM; AS
Aerobic-Primary
SM; AS; SL SM; AS SM; AS AS; WNV
Aerobic-Ultimate
SM; AS; SL SM; AS
Aerobic-Primary
SM; AS SM; AS
Aerobic-Ultimate
SM; AS (cont.) FW
Test Matrix
Initial Conc. (ppm)
28
28 1
28 28 28 21
28 1 9
28 27
3.5 5
Test Duration (Days)
Table 6: Summary ofPhthalate Biodegredation Literature
37
> 50 51
56 67 67 42
> 99 68 90
98 86
55
90
(%)
Degradation
6 55 26 15
6 42
3 42
Reference
OO
Fresh water (e.g., river, lake, pond) Landfill leachate Municipal Solid Waste Marine Water No / Yes Not Reported Synthetic Compost Sediment Sludge Soil Synthetic aqueous medium Wastewater
1. Barth and Bunch, 1979; 2. Urushigawa and Yonezawa, 1979; 3. O'Grady et al. 1985; 4. Tabak et al. 1981; 5. Patterson and Kodukula, 1981; 6. Suggat et al. ! 984; 7. Ritsema et al. 1989; 8. Zoetman et al. 1980; 9. Hattori et al. 1975; 10. Shanker et al. 1985; I 1. Russell et al. 1985; 12. Kinconnin and Lin, 1985; 13. Madsen et al. 1995; 14. Furtmann, 1993; 15. CITI, 1992; ! 6. Osipoff, 198 I. 17. Shelton and Tiedje, 1981 ; 18. Ziogou et al. 1989; 19. Aichinger et al. 1992; 20. Maag and Lokke, 1990; 21. O'Connor et al. 1989; 22. Horowitz et al. 1982; 23. Battersby and Wilson, ! 989; 24. Lewis et al. 1984; 25. Overeash et al. 1986; 26. Association of the Plasticizer Industry,Japan, 1994; 27. Sehouten et al. 1979; 28. Kodama and Takai, 1974; 29, Steen et ai. 1980; 30. Walker et ai. 1984; 31. Cripe et al. 1987; 32. Tagatz et al. 1986; 33. Oven,ash et al. 1982; 34. Shelton et al. 1984a; 35. Painter and Jones, 1990; 36. Snell Environmental Group, 1982; 37. Scholz, 1994a; 38. Johnson and Lulves, 1975; 39. Inman et al. 1984; 40. Johnson and Heitkamp, 1984; 41. Battersby and Wilson, 1988; 42. Saeger and Tucker, 1976; 43. Saeger and Tucker, 1973; 44. Gledhill et al. 1980; 45. Adams et al. i 989; 46. Adam and Saeger, 1993; 47. Sanborn et al. 1975; 48. Chaure~ et al., 1996; 49. Wooek, 1979; 50. Perez et al. 1985; 51. lrvine et al. 1993; 52. Schwartz et al., 1979; 53. Seholz, 1994b; 54. Struijs and Stolenkamp, 1990; 55. Exxon Biomedical Sciences, Inc., 1995a; 56. Davey et al. 1990; 57. Subba-Ran et al. 1982; SS. Wylie et al. 1982; 59. Rubin et al. 1982; 60. Sodergren, 1982; 61. Fairbanks et al. 1985; 62. Efroymson and Alexander, 1994; 63. Schmitzer et al. 1988; 64. Kloskowski et al. 1981; 65. Aranda et al. 1989; 66. Scheunert et al. 1987; 67. Fish et al. 1977; 68. Nyholm, 1990; 69. Exxon Biomedical Sciences, Inc., 1995b; 70. Exxon Biomedieal Sciences, Inc., 1996b; 71. Ye Changming and Tian Kang, 1990; 72. Rudel et al., 1993; 73. Kirehman et al., 1991; 74. Dorfler et al., 1996; 75. Ejlertsson et al., 1996.
References;
FW LL MSW MW N/Y NR SC SD SG SL SM W%V
Table 6: Summary of Phthalate Biodegradation Literature (cont.)
o~
700
I
I
I
I
I
DTE~ DUP
I
I
I
I
rl
0
O • ¢~ 711P
•
•
QD
DEeP
-J
0
COQ n
(3
~. el0P 0
BOP
<
mr
O r, "=l)-
BlIP
0 m• I
0
I
10 20
I
I
I
I
I
I
80
40
50
60
70
80
I-
I
O0 100
Ultimate Blodegradatlon
Figure 2. Ultimate biodegradation of phthalate esters in standardized tests based on CO2 evolution (circles) or 02 uptake (squares). Filled symbols indicate acclimated inoculum was used.
degradation within 28 days, with the exception of DTDP (Figure 2). The extent of ultimate degradation also tends to decrease as the size of the alkyl chain increases. Anaerobic inoculum tests have been conducted by Gledhill et al. (1980); Shelton and Tiedje (1981); Horowitz et al. (1982); Shelton et al. (1984); Shelton and Tiedj¢ (1984); Battersby and Wilson (1988, 1989); O'Connor et al. (1989); Ziogou et al. (1989); Painter and Jones (1990); and Madsen et al. (1995). Results of primary and ultimate biodegradation ~tudies under anaerobic conditions tend to be more variable than aerobic tests (Table 6). However, several generalizations can be made. First, the rate of degradation under anaerobic conditions is retarded relative to aerobic tests. Second, a similar, but less well-defined trend is observed between alkyl chain length and the extent of anaerobic degradation. Increased variability in anaerobic biodegradation
701 suggests that the nature of the inoculum influences test results. This is supported by the work of Horowitz et al. (1982) who found that the source of sludge had a pronounced effect on the rate of anaerobic biodegradation of selected phthalate esters.
Primary biodegradation half-lives of phthalate esters in natural waters were estimated from data in Table 6 assuming first order kinetics. For studies in which test duration or percent biodegradation were reported as ranges, average values were used to compute mean half-lives. For studies in which 100% degradation is reported, half-lives were estimated by dividing the test duration by three. Although such an approach is simplistic, this analysis provides a convenient method to normalize the extensive biodegradation database assembled. Data used to construct Figure 3 are based on observed biodegradation in unacclimated test systems, using fresh or marine waters, with or without sediment, under aerobic (or partially aerobic) conditions. Studies by Perez et al. (1985), Ritsema et al. (1989), and Furtmann (1993) suggest primary biodegradation rates are significantly retarded at low temperatures. Therefore, tests performed at low temperatures (e.g. < 5"C) were excluded in this analysis. Half-lives shown in Figure 3 range from <1 day to about 2 weeks. Clear trends between different phthalate esters or between freshwater and marine waters are not apparent. Limited data for DBP and BBP suggest that addition of sediment does not significantly influence primary biodegradation rates.
Research on the ultimate biodegradability of phthalate esters in natural waters has focused principally on DEI-IP. Using biodegradation test results reported by Wylie et al. (1982) with unfiltered water samples, a mean DEHP half-life of 78 days is calculated. Other studies by Subba-Rao et al. (1982) and Rubin et al. (1982) yield half-lives for DEI-IP ultimate biodegradation in eutrophic lake water ranging from 12 to 64 days. In contrast, no DEHP mineralization was observed after 60 days in water from an oligotrophie lake (Rubin et al. 1982). The lack of mineralization in the nutrient Ix)or lake water was attributed to either the absence of competent microbial populations or the lack of growth factors required by indigenous degraders. In a marine microcosm study by Davey et al. (1990), the ultimate degradation half-life for DEHP averaged 12 and 67 days, respectively at temperatures <10 and >20 C, respectively. Comparison ofth6~¢ results to primary biodegradation half-lives for DEHP (Figure 3) suggests ultimate biodegradation half-lives are about art order of magnitude higher.
Half-lives in soils for the more well studied phthalate esters were calculated using data in Table 6 and the same approach described previously (Figure 4). Data used for this analysis were based on unacclimated test systems that did not exclude oxygen, but in many cases, likely reflect only partially aerobic conditions. Results repgrted by Kloskowski et al. (1981) for DEHP were not used due to the short time frame of this experiment (7 days). Figure 4 includes soils half-lives which exhibit considerable variation ranging from about 1 to 75 days for primary and 4 to 250 days for ultimate biodegradation. Such variation likely reflects differences in the biodegradation test protocols employed and
702
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.
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.
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.
.
.
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o
o
o
ao
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o
ago
o
o
o
o ,
......
,
0.1
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,
, .....
1.0
,
,
,
,,
r-
10.0
100.0
Prlm~-y Degra,'~tlon I-mr-Life (,~,,ye)
Figure 3. Primary biodegradation half-lives of phtbolate esters in fresh (circles) and marine (squares) waters under aerobic conditions. Filled symbols denote results for test systems containing sediment.
D~'IP
~D
DSP
•
DEP
eO
Q
•
,
,
,
....
0
e6
•
DMP
0.1
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,
10.0
. . . . . . . .
,
1000
,
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1000.0
De~'amam H~a~Je emy~
Figure 4. Primary (open symbols) and ultimate (filled symbols) biodegradation half-lives of phthainte esters in sofia.
703 in the microbial populations associated with the various soils investigated.
Limited data are available on
biodegradation in sediments. Based on lab and field studies, Tagatz e( al. (1986) reported primary biodegrad~tlon rates in sedimems of 3=4 weeks for DBP. In an earlier lab study, Fish et al. (1977) provide primary biodegradation data indicating a half-life for sediment-bound DEHF of ca. 3 months. In an anaerobic test system, Madsen et al. 0995) found that ultimate biodegradation of DMP and DIBP added to either a freshwater swamp or marine sediment was <30% after 56 days. Johnson & Heitkamp (1984) showed that the ultimate biodegradatinn of DIOP and DINP in sediments was generally <10% after 28 days. These studies suggest that primary biodegradation half=lives are on the order of several months with ultimate biodegradation proceeding more slowly with half=lives that may exceed I00 days. Since oxygen may be limiting in soils/sediments, the longer half=lives observed in ihese matrices when contrasted to natural waters (Figure 3) are consistent with the limited extent of biodegradation observed in anaerobic inoculum tests discussed previously.
Several generalizations are apparent from our review of phthalate ester biodegradation literature. First, phthalam esters may be used by aerobic and anaerobic microbes as a source of carbon and energy. Second, phthalate esters are degraded in both standardized biodegradation tests and in tests simulating different environmental compartments. The extent of biodegradation occurring over the course of these tests suggests that phthalate esters are not expected to be highly persistent in most environments as is sometimes incorrectly asserted. Third, primary degradation in water, sediment and soil coropartments is expected to be controlled by biodegradation rather than abiotic loss mechanisms. Fourth, based on available data for unacclimated test systems, the primary degradation half-life in water is expected to be on the order of less than one week, while the half-lives in soils ranges from less than one week to severalmonths. These values are consistent with estimates obtained by several recent independent reviews on phtbalat~ ester bio~gradafion (Peijnenburg et al., 1991, Howard 1991; Boethling et al., 1995). Lastly, longer half-lives are likely under anaerobic conditions and in cold, nutrient poor environments.
BIOACCUMULATION
Bioaccumulation refers to the accumulation of contaminants in the tissue of biota via all exposure routes (i.e., water + diet) whereas bioconcenu'ation refers to accumulation due to aqueous exposure alone (Veith et al., 1979a). These processes are quantified by the bioaccumulation and bioconcentration factor, respectively (i.e., BAF or BCF). The BCF and BAF define the ratio of concentrations in tissue to that in water and am expressed in units of mL per gram body weight, on a dry, wet or lipid basis. The BAF for a chemical can be derived based on field data using measmed water and tissue levels or can be predicted using bioaccumulation models if sufficient toxicokinetic ~
are available
(Tbomann, 1989; Clark et al., 1990; Oobas, 1993). In contrast, the BCF is derived from a laboratory test and ¢xtaesses the partition coefficient of a chemical between test organism and water at steady-state. In most oases, bioc,omentmfion
704 can be adequately described by simple first-order kinetic models so that the BCF may be calculated by dividing the uptake clearance, K , in units ofmL g-t day-~,by the elimination rate, K2, in units of day ~.
Numerous laboratory studies have investigated the bioconcentration, and to a lesser extent, bioaccumulation of phthalate esters. Many investigators have used radiolabeled phthalate esters due to the greater sensitivity and convenience afforded by radioisotopes. For chemicals such as phthalate esters that are subject to metabolism, BCFs based on total radioactivity are expected to be higher than BCFs based on parent compound since radiolabeled metabolites can contribute significantly to total radioactivity. A comprehensive compilation of phthalate ester bioconcentration studies with fish and other aquatic organisms is provided in Tables 7 and 8, respectively.
In cases where toxicokinetic parameters were not reported but raw data were provided, the original data were reanalyzed using the non-linear regression procedure included in SYSTAT (Wilkinson, 1990) to estimate these parameters and the resulting steady-state BCF. Results provided in Tables 7 and 8 include several studies in which organisms were fed or had the opportunity to feed (e.g., microcosm studies) on contaminated prey. Therefore, some of the reported BCFs may be indicative of BAFs due to dietary exposure.
In the microcosm studies by Metcalf et al. (1973), Sanborn et al. (1975) and Perez et al. (1985), BCFs were estimated by dividing the total ~4Cradioactivity in the various test organisms at the end of the experiment (30-33 days) by the average )4C activity in the aqueous phase during the test. For the first two studies, this average was based on measurements at the start and end of tests. Since total activity in the water column decreased by less than a factor of four in both studies this average provides a reasonably constant exposure concentration for calculating BCFs. In contrast, total activity dropped three orders of magnitude in a 27-day DEHP microcosm experiment reported by Sondergren (1982). Due to the large variation in exposure concentration it was judged that reliable BCFs could not be determined from this study.
Severul early investigations that used chromatographic techniques to distinguish parent phthalate esters from )(C metabolites, raised environmental concerns about phthalate esters as a result of the high BCFs reported (Metcalf et al., 1973; Sanborn et al., 1975; Sodergren, 1982). However, these values are misleading because calculations are based on parent phthalate ester concentrations determined in tissue and water samples at test termination only. To provide more realistic BCFs for parent phthalate esters that are derived from radiotracer studies, the BCF, based on total n4C, was multiplied by the percentage of the total radioactivity that was reported in tissue as parent compound at test termination.
Studies with fish (Table 7) indicate that Kt and BCF increase with increased t~-mporatureand decreased body weight (Karara and Hayton, 1989; Tart et al., 1990). For the more hydrophobic phthalate esters, K. and BCF tend to decrease
NR (-2.5)
0.09 ± 0.04
Pimephalespromelas
(Fathead Minnow)
(Sheepshead Minnow)
Cyprinodon variegatus
(English Sole)
Parophrys vetulus
(Bluegill Sunfish)
28.3 4- 1.0
blR (-l.0)
Lepomis macrochirus
(Bluegill Sunfish)
NR (~1.0)
0.37 - 0.94
Lepomis macrochirus
(Bluegill Sunfish)
Lepomis macrochirus
(End,fish Sole)
Parophrys vetulus
(Bluegill Sunfish)
32.7 ± 1.6
0.37 - 0.94
Lepomismacrochirus
(Sheepshead Minnow)
NR (-2.5)
0.37- 0.94
Weight (g wet)
Cyrinodon variegatus
(Bluegill Sunfish)
Lepomis macrochirus
Organism
25.4
N-R
12
?
NR
16
i2
16
NR
16
Temp (°C)
FT
S
FT
FT?
FT?
FT
FT
FT
S
FT
Expos.
ad. lib.
Yes
No
No
?
?
30 -200
9.4
Diethyl P h t h a l a t e
I00
8.7
5 50
100
[62l
40 - 81
6
19 - 39
.
.
165 145
11.7
[400]
.
.
225 - 457
Di-n-Butyl P h t h a l a t e
20 - 250
34
2
9.7
K, ,
.
.
(mL/g wet'ld I)
Butylbenzyl P h t h a l a t e ad. lib.
No
ad. lib.
No
Conc. ~g/L)
Exposure
Dimethyl Phthalate
Feeding
Test Conditions
.
.
0.08 0.07
. . . .
. . . .
0.34 - 0.69
. . . .
0.34 - 0.69
2,068 2,125
4494 453
293
1882
663
117
57
167 s 172 s
__
I
12
Total Parent BCF I BCF' (mL/g wet "l) (mL/g wet")
. . . .
0.34 - 0.69
(d "l)
K2 I
Table 7. Summary o f Bioconcentration Studies for Phthalate Esters with Fish.
Ref.
-,O C2~
! .24 ±
0.31 1.24 ±
0.31 1.24 ±
0.31 1.24 ±
0.31 ! .24 ±
0.31 1.24 ±
0.31
Pimcphalespromelas
(Fathead Minnow)
Pimephalespromelas
(Fathead Minnow)
Pimephalespromelas
(Fathead Minnow)
Pimephalespromelas
(Fathead Minnow)
Pimephalespromelas
(Fathead Minnow)
Pimephalespromelas
(Fathead Minnow)
(Blue[gill Sunfish)
Lepomis macrochirus
1.0 - 2.0
0.31
(Fathead Minnow)
22?
25
25
25
25
25
25
FT
FT
FT
FT
FT
FT
FT
FT
! .24 +
Pimephalespromelas
Gambusia sp. 25
S
26
NR (- 0.1)
(Mosquito Fish)
(carp)
FT
Expos.
?
Temp (*C)
Exposure Cone. 0zg/L) K, J (mL/g wet"d')
25
5% bdw
Yes
Yes
Yes
Yes
Yes
Yes
Yes
N Y
5.7 5.6
62
30
14
8.1
4.6
2.5
t.9
100:1: 19
--
767
8.2 [21
17 [6]
76 [9]
132 [161
80 [34]
82 [69]
707 [29] 9
2046
Di(2-Ethylhexyl) Phthalate
Yes
D i - n - B u t y l P h t h a l a t e (cont.)
Feeding
Test Conditions
?
Weight (g wet)
Cyprinus carpio
Organism
0.67 ~ 0.47
0.05, 0.04 I0.041
0.06, 0.02 [0.05]
0.25, 0.03 [0.05]
0.34, 0.05 [0.06]
137 + 6 I°
114+87
155 + 36
295 ± 65
304 + 51
384 + 25
817±72
900 + 69 0.10, 0.06 [0.06]
745 + 627
[0.05]' 0.09, 0.06 [0.11 ]
11
51
124
182
270
564
630
588
2
3.6
Total Parent BCF L BCF I (mL/g wet") (mL/g wet")
0.097, 0.048
--
K2' (d")
Table 7. Summary of Bioeoncentration Studies for Phthalate Esters with Fish.
lO
9
18
Ref.
Oncorhynehus mykiss,
(Carp)
Cyprinus carpio
(Mullet)
Mugil cephalus
formerly Salmo gairdneri (Rainbow Trout)
Oncorhynchus mykiss,
formerly Saimo gairdneri (Rainbow Trout)
30
4.3 :t: 0.4
2.9 61.3 441
25
24
12 12 12
I0 lO 10
NR (--O.1) NR
(-0.O
l0 16 23 29 35
NR
20-25
12
16
Temp (°C)
-5 -5 -5 - 5 - 5
1 1 1 1 !
Cyprinodon variegatus
(Sheepshead Minnow)
(Sheepshead Minnow)
NR (-2.5)
I -5
26.7 ± 1.3
0.37 - 0.94
Weight (g wet)
Cyprinodon variegatus
(Golden orfe)
Leuciscus idus melanotus
(English Sole)
Parophrys vetulus
(Bluegill Sunfish)
Lepomis macrochirus
Organism
FT
FT
S S S
FT FT FT
S S S S S
S
S
FT
FT
Expos.
Conc. (#g/L)
Exposure
No No
2% bdw
2 - 4% bdw
No No No
0.1-1.0
50
30 30 30
5 14 54
60
60 60
No
5% bdw 5%bdw 5% bdw
60 60
! 00 500
50
25 - 300
5.8
No No
No
No
No
ad. lib.
. . .
85 - 104
0.39
[2.3]
[3.71
[1.25]
[l.S]
. . .
[0.7] [0. l 1]
[0.5]
[I .0] [2.3]
.
0.23
K2] (d 1)
[16.1]
[64.6]
. . .
[672] [720]
[317]
[46] [250]
! 0.7 13.5
[24]
26
. . .
, Ki (mL/g werld ')
D i ( 2 - E t h y l h e x y l ) P h t h a l a t e (,:ont.)
Feeding
Test Conditions
.
.
Total
-
--
--
--
78 113 42
---
--
-
.
220 - 270
.
40'
114
BCF j
.
1.0 - 29.7
51.5 8.9 1.6
15 19 16
962 u 6,510 u
637
6 !3!
Parent BCF I (mL/g wet "]) (mL/g w e r ' )
Table 7. Summary o f Bioconcentration Studies for Phthalate Esters with Fish.
16
15
14 14 14
13 13 13
12,17 12,17 12,17 12,17 12,17
2
II
3
I
Ref.
¢>
25
12
26
Temp (°C)
FT
FT
S
Expos.
30 - 40
6.4
O. I - 1.0
.
[70]
.
.
.
.
K it (mL/g wetld I)
Diisodecyl Phthalate 2% bdw
No
Y
Cone. ~g/L)
Exposure
Diisooctyl Phthalate
Feeding
Test Conditions
.
.
.
.
.
K~ (d -I)
.
.
Total
Parent
.
.
.
207 .
<3.6 - < 14.4
92
BCF ~ BCF ~ (mL/g wet ~) (mL/g wet -~)
t6
3
7
Ref.
References: 1, Barrows et al. 1980; 2. Wofford et al. 1981; 3. Boese, 1984; 4. Heidolph and Gledhill, 1979; 5. Carr et al. 1992; 6. Call et al. 1983; 7. Sanbom et al. 1975; 8. Metcalfet al. 1973; 9. Mayer, 1976; 10. Macek et al. 1979; 11. Freitag et al. 1985; 12. Karara and Hayton, 1989; 13. Mehrle and Mayer, 1976; 14. Tart et al. 1990; 15. Park et al. 1990; 16. CITI, 1992; 17. Karara and Hayton, 1984; 18. Scholtz and Diefenbach, 1996.
S = Static; FT = Flow Through; NR = Not Reported; % bdw = Percent of Body Weight Fed Daily; ad. lib. = ad libetum; [ ] Denotes values that are based on parent compound; :[: = Nominal concentration at start of test.
Footnotes Values reported are based on total radioactivity determined in whole body and water and do not distinguish parent compound from metabolites unless otherwise indicated. 2 BCF after 17 day exposure. 3 BCF for fillet. 4 BCF after 3 day exposure. s Calculated using fraction of total radioactivity reported as parent compound after 11 days. 6 Calculated from reported tissue concentration after 1 hour exposure. 7 Calculated from data obtained during uptake phase.. s Calculated from data obtained during deputation phase. 9 Calculated by multiplying BCF for parent compound by the i st order elimination rate of parent compound from deputation phase. ~0 Includes dietary exposure in which daphnids containing 28 IJg/g DEHP were fed to fish at a 5% bdw ration. '~ BCF test performed under elevated temperatures (29-35 C).
(Carp)
Cyprinus carpio 30
38.5 + 2.7
Parophrys vetulus
(English Sole)
NR ( - 0.1)
Weight (g wet)
(Mosquito Fish)
Gambusia sp.
Organism
Table 7. Summary of Bioconcentration Studies for Phthalate Esters with Fish.
oo
Crustacea
Algae
Mollusca
lnsecta
30 21 21 21 2!
Artemia
Daphnia magna
Hexagenia bilineata
Ischnura verticalis
Chironom~
FT FT
FT
21 2! 21
21
Selenastrum
Daphnia magna Daphnia magna Daphnia magna
S
24
S
SR
SR
SR
SR
S
FT
Physa
phAmosz~s
! O-12
SR
21
Gammarus pulex
pseudolimnaeus
Gammarus
Palaemontes kadiakensis
Cmstacea
SR
S
HR 21
S
NR
Selenastrum
Algae
S
Expos.
30
Temp. (°C)
Test Conditions
Pseudomonas
Organism
Bacteria
Taxa
.
2700-13,900 0.08 0.08 0.1 0.18
No No No No No
0.08 0.21 0.21 0.21
No Yes Yes
Dihexyl P h t h a l a t e
1.4
NA
No
Butylbenzyl Phthalate
100~:
Yes
128 129 371
.
33
350
133
99
189
.
.
416
0.1
NO
248
0.08
.
No
2,000 10,000
NA NA
1,000 - 7,000
NA
.
.
.
.
.
.
.
.
.
.
.
.
.
KI I (mL g weted "l)
Di-n-Butyl P h t h a l a t e
Feeding
Exposure Cone. (/zg/L)
.
.
.
.
.
.
. 0.13 0.09 0.07
.
0.33
0.5
0.29
0.14
0.47
0.28
NC
.
.
.
K21 (d "l)
.
.
.
. 999 !,414 s 5,2699
100
700
4583
714
403
3453
1857
1485
7503.6
( m L g wet "l)
Total BCFI
Table 8. S u m m a r y o f Bioconcentration Studies for Phthalate Esters with Other Aquatic Organisms.
----
1,3752
--
--
--
--
--
--
__
--
--
1,324
5,4752'3
78 - ! 802'3
( m L g wet "N)
Parent BCFI
8 8 8
8
7
4
3
4
4
6
5
4
3
2
1
Ref.
26
26 20-25
Daphnia magna
Culexpipiens
Dedogonium
Chlorella
Crustacea
Insccta
Algae
26
26 26 24 15 15 1° I° 1° 18 ° 18 °
Elodea
Physa
Physa
Physa
Mytilus edilus Mytilus edilus
Pitar morrhuana Pitar morrhuana Pitar morrhuana Pitar morrhuana Pitar morrhuana
Aquatic Maerophytes
Mollusca
26
26
26
ehysa
Algae
FT FT FT FT FT
FT FT
S
S
S
S
S
S
S
S
S
S
Expos.
Test Conditions Temp. (°C)
Mollusca
Organism
Dedogonium
Taxa
6.4
6.4
6.4
6.4
Yes Yes Yes Yes Yes
Yes Yes
No
Yes
No
NA
NA
NA
0.6 5.9 59 0.2 1.2
4 42
0.35
19
100
100
50
19
K~~
473 525
76
.
73
.
.
.
.
.
.
.
.
.
.
.
.
.
.
(mL g wet-ld")
Di (2-Ethylhexyl) Phthalate
Yes
Yes
Yes
NA
Cone. (,ug/L)
Di-n-Octyl Phthalate
Feeding
Exposure
.
.
.
.
.
.°
-.
0.20 0.20
0.29
0.09
--
(d-I)
K21
2,366 2,627
264
1,050
814
987
1,338
1,429
699
8,412
(mL g wet"l)
Total BCFt
Parent BCF I
1,3805 1,047 436 1,122 363
377
811
2754
5,4002
946
91
26
133
278
(mL g wet-I)
Table 8. Summary of Bioconeentration Studies for Phthalate Esters with Other Aquatic Organisms. (cont.)
13 13 13 13 13
12 12
7
10
10
10
II
10
9
9
9
9
Ref.
"-,.I
Cmstacea
Mollusca (cont.)
Taxa
D
10- 12 25 25 NR
Gammaruspulex
Asellus brevieaudus
26
22 20 20 20
Daphnia magna
Daphnia magna
Daphnia magna
20
21
Daphnia magna
vannamei
Panaeus
SR SR
21 2!
Gammarus pseudolimnaeus
SR SR SR SR
S
S
SR
FT
SR SR
FT FT FT FT FT FT
l* 1* 1* 18* 18 * 18*
Nucula annulata Nucula annulata Nucula annulata Nucula annulata Nucula annulata Nucula annulata
FT F'I" FT FT FT FT
Expos.
1* 1* 1* ! 8* 18" ! 8"
Temp. (*C)
Feeding
Cone. 0zg/L)
Exposure
Yes Yes Yes Yes
No
No
No
Yes
No No
Yes
No No
Yes Yes Yes Yes Yes Yes
Yes Yes Yes Yes Yes Yes
2.2 7.3 25.3 85.6
5.5
100:[:
0.3
60 - 6,000"
1.9 62.3
100
0.10 62.8
0.6 5.9 59 0.2 1.2 15.5
0.6 5.9 59 0.2 1.2 ! 5.5
. . . -
.
NC
61
--
37 NC
.
822 8.3
. . . . . .
. . . . . .
. . .
.
.
. . . . . .
. . . .
. .
. . .
.
.
. . . . . .
. . . .
.
(mL g wetld a)
K~t
Di (2-Ethylhexyl) Phthalate (cont.)
Test Conditions
Mulinia lateralis Mulinia lateralis Mulinia lateralis Mulinia lateralis Mulinia lateralis Mulinia lateralis
Organism
. . .
.
.
.
.
. . . . . .
. . . .
.
--
NC
NC
0.81
3.4 NC
0.21 0.10
. . . . . .
. . . .
.
K2J (da)
.
. . . . . .
. . . .
.
166 140 261 268
518-'
1834
41612
--
113 383'~°
14907
3916 83
Total BCF I (mL g wet"l)
-
-
179
1,995
1,820
174 100 1,9O5
16 16 16 16
15
10
4
14
3 3
5
3
4
13 13 13 13 13 13
! 26 s
933 1,175 3,31 I
3,890
13 13 13 13 13 13
Ref.
2,455 s 3,090
Paint BCF I (mL g wet "m)
Table 8. Summary o f Bioconeentration Studies for Phthalate Esters with Other Aquatic Organisms. (cont.)
21
21 22 26 5°
26.5 26.5
15 15 20 20 20 20
Hexagenia bilineata
Chironomus plumosus
Culex pipiens
Rana arualis
Diplora strigosa
Arca zebra
Mytilus edilus Mytilus edilus
Daphnia magna Daphnia magna Daphnia magna Daphnia magna
Insecta
Amphibia
Coelenterata
Mollusea
Mollusea
Crustacea
18 o 18 °
1o 1° 1°
NR
incisa incisa incisa incisa incisa incisa
SR SR SR SR
FT FT
S
S
S
S
SR SR
SR
D
FT FT FT FT FT FT
Expos.
Feeding
Exposure Conc. (/zg/L)
Yes Yes Yes Yes
Yes Yes
No
No
2.9 9.6 32.6 100
4.4 41.7
Diisodeeyl Phthalate
61
61
Phthalate
0.9 - 187
19
0.3 0.2
0.1
Diisononyl
Yes
Yes
No NR
No
10.05
5.9 59 0.2 1.2 15.5
Yes Yes Yes Yes Yes Yes
0.6
Yes
. . . .
795 539
848
--
.
.
568 81.6
128
--
. . . . .
.
. . . .
. .
. . . . .
.
. . . .
. .
. . . . .
.
K~~ (mL g wet-ld -I)
Di (2-Ethylhexyl) Phthalate (cont.)
Test Conditions Temp. (°C)
Lumbriculus terrestris
Nepthys Nepthys Nepthys Nepthys Nepthys Nepthys
Organism
Annelida
Polychaeta
Taxa
. . . .
.
. . . . .
.
.
0.20 0.18
0.46
0.02
.
1.8 0.2
0.14
0.04
. . . . .
.
K2 t (d "l)
.
. . . . .
.
100 147 128 90
3,977 2,998
1,844
--
1,892
315 408
915
.
.
.
Total BCFI (mL g wet -I)
- .
- .
. -
16 16 16 16
12 12
20
20
19
450 - 76013
4 18 10
- -
4
17
13 13 13 13 13
13
Ref.
1,892
.
47 30 I,I 22 724 549
595
Parent BCFI (mL g wet "l)
Table 8. Summary o f Bioconeentmtion Studies for Phthalatc Esters with Other Aquatic Organisms. (cont.)
I O
References: 1. Wang and Gradie, 1994; 2. Casserly et al. 1983; 3. Sanders et al. 1973; 4. Mayer and Sanders, 1973; ft. Thuren and Woin, 1988; 6. Hudson et al. 1981; 7. Parkerton, 1993; 8. Gloss and Biddinger, 1985; 9. Sanborn et al. 1975; 10. Metcalf et al. 1973; I 1. Geyer et al. 1981; 12. Brown and Thompson, 1982b; 13. Perez et al. 1985; 14. Hobson et al. 1984; 15. Macek et al. 1979; 16. Brown and Thompson, 1982a; 17. Albro et al. 1993; 18. Streufert et al. 1980; 19. Larsson and Thmen, 1987; 2O. Solbakken et al. 1985.
NR = not reported; NC ---not calculable due to erratic nature of reported tissue concentration during uptake phase; FT = flow through exposure; SR = static renewal exposure; S = static exposure; D = dietary exposure; 1: = nominal concentration at start o f test
Table 8. Summary of Bioconcentration Studies for Phthalate Esters with Other Aquatic Organisms. (cont.) Footnotes: = Values reported are based on total radioactivity determined in whole body and water and do not distinguish parent compound from metabolites unless= otherwise indicated. 2 BCF after I day. 3 Literature value converted from day to wet weight basis by assuming a dry/wet weight ratio of 0. ! 5. 4 BCF after 2 days. 5 BCF after 30 days in marine microcosm; calculated values include potential dietary exposure. 6 BCF after 3 days; steady-state not reached. BCF after 10 days based on the sum o f reported absorbed and adsorbed tissue concentrations. s Includes dietary exposure in which algae containing 0. I 1 ttg/g were fed to daphnia at a 1% bdw ratio. 9 Includes potential dietary exposure in which uncontaminated algae were fed to daphnia at a I% bdw ratio. io BCF after 21 days. is Concentration (ppm) administered in diet at 1% bdw ration for 4 days. 12 BCF after 21 days. J3 BCF after 60 days.
t.J
..,O
714 ~5th increased exposure concentration (Mayer, 1976). Boese (1984) hypothesized that this inverse relationship was likely due to test concentrations exceeding water solubility since the bioavailability of colloidal/particulate chemical species are expected to be reduced. Elevated temperature appears to reduce the elimination rate and significantly increase the overall BCF (Karara and Hayton, 1989). This observation may be due to the inhibition of temperature-sensitive enzymes that mediate biotransformation of phthalate esters. Application of chemical inhibitors that block enzymatic hydrolysis of diesters has also been shown to enhance phthalate ester bioconcentration in fish (Melancon et al., 1977; Karara and Haytorl, 1988).
In an attempt to compare phthalate ester toxicokinetic data for fish with results for other chemical classes, bioenergetics models were used to estimate K~ and K,_. K~ was calculated as described by Thomalm (1989): R
K] :
(3)
where: KI
= Uptake clearance [mL g wet t day ~]
R
= Respiration rate [mg 02 g wet" day ~]
E
= Efficiency of chemical uptake relative to that of oxygen [unitless]
Cox
= Dissolved oxygen concentration [mg 02 mL ~]
Respiration was estimated from fish weight (W, g wet) and experimental temperature (T, °C) using the empirical equation presented by Thurston and Goerke (1987): R = 1.32W -°]gs T °'29 exp °'°3s r
(4)
The ratio of the chemical to oxygen uptake efficiency (E) for fish was estimated based on the literature review by Parkerton (1993) for nonionic organic chemicals: For Log K.,, < 4; E = 1 - 0.3 (4 - Log Ko~,)
(5)
For Log Ko. ~ 4 but ~ 6, E :
1
(6)
and for Log Ko~ > 6; E
= 1 - 0.3
(Log
K,,,
-
6)
(7)
715 For these calculations, the dissolved oxygen concentration at saturation for the prescribed experimental conditions was assumed. For several studies included in Table 7, fish weights were not reported so reasonable values we're assumed based on literature data. The elimination rate was estimated according to Thomann (1989): Ki = - -
(S)
where f,pis the lipid fraction of the fish.A typical ft~pof 0.05 was selected for calculations. A comparison of predicted and observed phthalate ester toxicokinetic parameters for fish is illustratedin Figure 5. The upper part of this figure indicates that the predicted uptake clearances systematically overestimate observed values. This bias may be attributed to presystemic metabolism of phthalate esters by esterases associated with the fish gill (Barren et al., 198g). As previously mentioned, a second explanation contributing to the lower observed KI values may be bioavailabilitylimitationsassociated with performing bioconcentration testsat concentrations in excess of water solubility.However, this rationale would only apply to the higher molecular weight phthalate esters.
Examination of the lower panel in Figure 5 shows a similar comparison for elimination rates. The predicted elimination rates overestimate data in the low Ko~ range (i.e., DMP, DEP) which is likely due to limits placed on chemical diffusion rates from aqueous fish compartments. In contrast, predicted elimination rates grossly underestimate observed rates for higher molecular weight phthalate esters since metabolism is ignored. The difference between the predicted and observed elimination rates can in fact be used to provide quantitative estimates of biotransformation rates (Sijm, 1992; DeWolfe et al., 1992).
Experimental fish BCFs from Table 7 are summarized graphically in Figure 6 as a function of log K~. The observed BCFs do not appear to be well correlated to Kow and tend to cluster around a value of 100. Comparison with predicted BCFs calculated from equations (3) and (8), and denoted by the solid line, indicates the model overestimates observed values by
approximately three orders of magnitude for the higher molecular weight phtha[atc esters (Figure 6). Again, these results are attributedto metabolism (Hogan, 1977; Metcalfet al.,1973; Carl 1992; Ban'on et al.,1995). In con~st, BCFs for the more water soluble phthalatc esters arc underestimated by about an order of magnitude due to the erroneously high elimination mtc that is predicted from equation (5). Data for other aquatic organisms also indicate BCFs are lower than expected based on simple lipid-water partitioning and do not correlate with phthalate ester hydrophobicity (Table 8).
For studies in which BCFs are reported in terms of both total '4C and parent phthalate ester, results are highly variable. BCFs for parent phth~ate esters range from less than a factor of two to as much as a factor of fiRy lower than BCFs
det=nnlned from totalradioactivity(Tables 7 and g). To provide a furthersynthesis of available BCF data,the average BCF for various taxa were calculated for the two most commonly studied phthalate eaters, DnBP and DEHP (Table 9). The average taxon BCF was determined by first calculating the mean BCF for each species within a taxa and then computing
716
1000
(a) . . . . . . . . . . . . . . . .
•
0-.
o~"o
u
~
°°°
~a
°
J
'
°
' " ~
•
v/~:l
/
"
"°" °o
°
0 DMP DBP t '~vBBP
I
100
0 ~P ~DIOP
0
10
i
,
i
,
,,l,I
,
a
i
10
. . . . .
I
. . . . . . .
100
1000
Observed Uptake Olearanoe (mL g"a"~)
100.00000 10,00000
li
1.00000
#. 0,10000
DEHP D~P
0.01000 0.0010O
.,t
8~
0.00010 0.0O001 0,001
ee °
%
. . . . . . . .
i
0,010
. . . . . . . .
i
0.100
. . . . . . . .
i
. . . . . .
1.000
10.000
Obeerv~J Bl~natlon Rate (d"~)
Figure 5. Comparison of predicted and observed uptake clearances (a) and elimination rates Co) of phthalate esters in fish based on total radioactivity (open'symbols) or parent compound (filled symbols).
717
7
I
I
I
I
I
I
0 DMP ADi~ oDBP vBBP ODEHP ODIOP
6
4 8
1 I
I
I
i
I
I
2
8
4
5
6
7
8
Log Kow
Figure 6. Comparison of measured fish bioconcentration factors based on total radioactivity (open symbols) or parent compound (filled symbols) with model predictions (Hne) as a function of the octanol-water partition coeffident (Log K~).
718 Table 9. Summary of Mean Bioconcentration Factors for DBP and DEHP in Different "faxa.
Phthalate
Taxon
# Species
Mean BCF (ml g wet 1)
Di-n-Butyl
Algae
1
3399
Phthalate
Crustaceans
3
662 ± 229
Insects
3
624 ~- 144
Fish
1
167
Di (2-Ethylhexyl)
Algae
2
3173 + 3149
Phthalate
Molluscs
5
1469 ± 949
Crustacea
4
1164 + 1182
Insects
3
1058 ± 772
Polychaetes
1
422
Fish
5
280 ± 230
Amphibians
1
605
+ standard deviation of species mean BCFs
the mean of the species BCFs for that taxa. In cases where BCFs for parent phthalate ester were provided, these values were used in calculations. Otherwise, BCFs based on total ~4C can be used to provide conservative BCFs for parent phthalate ester. Results indicate mean BCFs are highest for algae and lowest for fish with invertebrates exhibiting intermediate values. These findings are consistent with studies by Wofford et al. (1981) who found that the extent of phthalate ester biotransformation increased as follows: molluscs < crustaceans < fish. It is interesting to note that the lowest elimination rate reported in Table 8 was observed for the marine coral,
Diplora strigosa, which is consistent with the more primitive
enzymatic system expected for coelenterates.
The specific metabolic pathway of phthalate esters in aquatic organisms has been investigated in a number of studies and reviews (Stalling et al., 1973; Johnson et al., 1977; Melancon, 1979; Lech and Melancon, 1980; Barton et al., 1995). The first step in the metabolism of phthalate esters is hydrolysis to the monoester and the corresponding alcohol. The alcohol is further metabolized via a carboxylic intermediate to acetate and carbon dioxide by [3-oxidation. The monoester can be: directly excreted; undergo Phase I reactions in which the alkyl chain is subjected to various oxidations via microsomal P450s and mitochondrial enzymes; or undergo Phase II reactions in which monoester is conjugated with glucuronide (Barton et al., 1995). Phthalate ester metabolism appears to depend upon both species and exposure route. Earthworms
719 lppear to be somewhat unique in lacking the capacity to oxidize monoester which may account for the relatively low ,-limination rate observed for this organism (Table 8).
Although numerous studies have investigated the bioconcentration of phthalate esters, comparatively little work has been reported on potential uptake through food. Dietary accumulation is usually described as a simple first order process (Bruggeman, 1981). At steady state, the concentration in a predator is given by:
c~,=Nc K2 prey
(9)
Where •
Cp,,,Lp,y = chemical concentration in predator or prey (diet) [~g g wet ~] I
= ingestion rate [g wet prey g wet predator'ld"~]
(X
= assimilation efficiency of dietary contaminant, and
K2
= the elimination rate defined previously.
Ma~k et al. (1979) examined dietary transfer to bluegills that were fed DFJ-IP contaminated daphnids while Hobson et ai. (1984) examined dietary uptake of DEHP by penseid shrimp fed a synthetic diet. Application of equation (9) to results based on total ~4C measurements obtained in these studies yield assimilation efficiencies of ca. 0.25-0.30. Assimilation efficiencies for pment DEHP may be lower as a consequence of metabolism in the gut. Gloss and Biddinger (1985) studied the importance of dietary transfer in daplmids fed DHP contaminated algae, but equivocal results preclude estimation of reliable assimilation efficiencies. In one set of experiments, accum,laHon from food+weter was 39% higher than water only exposures. However, inconsistent results were obtained in food-only experiments since no accumulation was observed. Perez et al. (1985) suggested that dietary exposure could account for seasonal ~
in the accumulation of DEHP by
marine biota in microcosm studies. Bivalve filter feeders had tenfold higher tissue concentrations in winter compared to summer experiments despite only a threefold increase in average DEHP water concentrations. It was suggested that the higher tissue concentrations could be explained by higher algal densities and subseqtamt ingestion that ocom'red during the winter or that lower winter temperatures decreased enzymatic activity. In contrast, benthic species had higher tissue concentrations in the summer that were attributed to elevated sediment ingestion rates associated with warmer ~
.
The relative importance of dietary and aqueous exposure can be determined by comparing the ratio of dietary and aqueous uptake:
¢~'~ = --~-~BAF
K I C,w ,
K,
(10)
720 To illustrate the relative importance of these mutes for DEHP, representative values for the parameters included in equation (10) can be selected for a 0.5 kilogram fish that feeds on invertebrates. An assimilation efficiency of 0.3 and an uptake clearance of 4 ml g wet1 day~ were selected based on laboratory studies (Macek et al., 1979; Tarr et al., 1990). A prey BAF of 2000 ml g wet"~was assumed based on the microcosm studies by Perez et al. (1985) while an ingestion rate of 0.01 g wet prey g wet predator "t day "1was obtained from the allometric equations suggested by Thomann (1989). Substituting these estimates into equation (10) yields a value of 1.5 indicating that as much as 60% of the DEHP exposure could be derived from the diet. In contrast, if the diet was comprised of smaller fish rather than invertebrates the prey BAF would be an order of magnitude lower and the dietary route would account for less than 15% of the DEHP accumulated. These results differ from estimates obtained for persistent organochlorine compounds with a Kow similar to DEHP which indicate more than 95% of the contaminant accumulation is due to the dietary route (Thomann, 1989). Consequently, increased metabolism and reduced prey concentrations between trophic levels prevents biomagnification of phthalate esters. In fact, phthalate ester concentrations are expected to decrease as one proceeds up the foodchain. This inverse pattern has been observed for other classes of metabolizable organic chemicals such as polyaromatic hydrocarbons (McElroy et al., 1989).
Bioaccumulation of sediment-associated phthalate esters by dragonfly and chironomid larvae has been investigated (Woin and Larsson, 1987; Brown et al., 1996). Assuming a typical lipid content of 1% wet weight for benthic invertebrates, biota to sediment accumulation factors (BSAFs) can be calculated. The BSAF expresses the ratio of the lipid-normalized biota concentration at steady-state to the corresponding organic carbon- normalized sediment concentration to which test organisms are exposed. The BSAF for DEHP in dragonflies is approximately 0.1 while the BSAF for both DEHP and DIDP in chironomids is about 0.5. These values likely overestimate BSAFs for parent compound since calculations are based on total 14C.Nevertheless, these BSAFs are lower than predicted from equilibrium partitioning theory (Lake et al., 1990) further supporting the importance of phthalate ester metabolism.
Imlma~_Qzumi.tz~ Bioaccumulation of phthalates by terrestrial plants has been the subject of several studies. Plant cuticle lipid-water partition coefficients were determined for three plants with a range of organic chemicals including DEHP and were found to be approximately equal to K ~ (Kerler and Schonherr, 1988). However, these studies employed radiolabeled compounds and did not determine if parent compound was accumulated. Shea et al. (1982) reported the uptake of DnBP contaminated soil by com. Plant to soil bioconcentration factors, defined as the ratio of wet weight concentrations in plant and soil, respectively, were less than 0.002 indicating the limited bioaccumulation potential of DnBP. Kirchmann and Tengsved (1991) investigated the uptake of DBP and DEHP in barley grown on soil fertilized with sludge containing these compounds. Application of sludge containing 37 mg/kg of DBP and 116 rdg/kg of DEHP to field plots at a rate of 5 tom/ha resulted in concentrations in barley at harvest of 0.116 mg/kg and 0.530 mg/kg dry, respectively. Concentrations
721 in plants harvested from control plots receiving no sludge addition were < 2 and ca. 5 time lower for DBP and DEHP, respectively. The authors concluded that uptake percentages of added phthalates by grains was very low amounting to only 0.1-0.2 % of the amount added to the soil. Overcash et al. (1986) investigated the uptake of t4C-DEP, DnBP, DnOP and DEHP from soil by corn, soybean, wheat and rescue. Plant species were grown to maturity in greenhouse pots in soil containingthree different treatment concentrations of each phthalate ranging from 0.02 to 4 mg/kg. Total ~4Cin plants at harvest indicated limited plant uptake with plant bioconcentration factors, expressed as the ratio of the plant concentration (wet weight) to the initial soil concentration (dry weight), ranging from <0.001 to 1.0 with most values < 0.1 and typical values <0.01 (Overcash et al., 1986). Since these values are based on total '4C, these values overestimate the true extent to which parent phthalates are bioconcentrated as supported by the findings of additional studies described below.
Uptake of )4C-DEHP by four species of plants (i.e., lettuce, carrot, chili peppers and tall rescue) grown on soil amended with DEHP contaminated municipal sludge was investigated by Aranda et al. (1989). Soil DEI-IP concentrations ranged from 2.6 to 14.1 mg/kg dry in these studies. Since parent DEHP was not detected in any of the plants grown on these soils, the authors concluded that plant uptake of DEHP was of minor importance and would not limit sludge application to soils used to grow the crops studied. These findings are consistent with laboratory and field studies reported by Schmitzer et al. (1988) who found that parent DEHP was not detected in barley and potatoes grown in soils contaminated with DEHP at concentrations ranging from 0.2 to 3.3 mg/kg. The low levels of total ~'C detected in plants from both studies were attributed to uptake of ~4COz produced from the biodegradation of ~4C DEHP during the course of the study. This explanation is further supported by the results of Kloskowski et al. (1981) who observed the mineralization of 14C DEI-IP in soil-plant systems. Schmitzer et al. (1988) cites earlier studies by Kato et al. (1980) who also reported DEHP was not accumulated by vegetables grown in DEHP contaminated soils. Foliar uptake of DBP and DEHP in air by plants also appears to be limited (Lokke and Bro-Ramussen, 1981, and Lokke and Rasmussen, 1983).
Data on the bioaccumulation of phthalates in terrestrial animals is limited but, like aquatic foodchalns, trophic transfer is expected to be of low concern as a result of extensive biotransformation. In one study, a simple detrital foodchaln consisting of woodlice feeding on fallen,4cer leaves contaminated with DEHP was used to investigate bioaccumulation potential in the tenestrial environment. Results indicated no DEHP bioconcenUation in adult woodlice or offspring during the 6 month microcosm experiment and showed that DEHP is degraded effectively in this terrestrial foodchain (Lokke, 1988). Several studies have investigated the bioaccumulation of selected phthalate esters in avian species. Belise et al. (1975) found that DBP and DEHP did not accumulate in tissues of mallard ducks above the limit of detection (0.15 mg/kg) after being fed diets comaining 10 mg/kg of either chemical for 5 months. Captive starlings were fed diets containing 25 and 250 mg/kg of DHP or DEHP for 30 days by O'Shea and Stafford (1980). Residue analyses at test termination indicated that DHP was not detected ( < 1 mg/kg) in any of the eight carcasses examined from either dietary dose. Five out of the eight birds contained detectable levels of DEHP. The average residue in samples with detectable concentrations was 1.8 mg/k8. In
722 another study, Ishiuda et al. (1982) fed white leghorn hens contaminated diets containing 50,000 and 100,000 mg/kg of DEHP for 230 days.
Various tissues were analyzed in two hens from each treatment at test termination. DEHP
concentrations in various muscles tissues ranged from 1.8-9.4 and 2.6-16.4 mg/kg for the 50,000 and 100,000 mg/kg dietary concentrations, respectively. In a subsequent 25 day feeding study in which hens were fed diets containing 2000 mg/kg of DEHP, muscle tissue concentrations ranged from <0.5 to 1.5 mg/kg. In both experiments, higher concentrations were observed in adipose tissues and feathers than muscle tissues. The elevated concentrations in feathers likely resulted from external contamination of the birds from the diet during feeding. Based on the results of the~ studies it is apparent that phthalate esters exhibit a very low bioaccumulation potential in avian wildlife.
Predictions that rely on simple correlations with chemical hydrophobicity for estimating bioaccumulation in agricultural foodchains such as proposed by Travis and Arms (1988) are inappropriate if applied to phthaiates since these models arc based on data for persistent chemicals. In a recent critique of the application of this model for assessing potential indirect human exposure to DEHP, Lowenbach et al. (1995) suggested an alternative approach to conservatively estimate bioaecumulation in meat and dairy products derived from contaminated areas. This approach relies upon an empirical correlation derived between bioconcentration in rodents and bioconcentration in earle for a large set of chemicals. Based on a measured rodent BCF for DEHP, the cattle BCF was estimated to be 0.008. This BCF represents the ratio of DEI-IP concentrations in cattle to that in the cattle's food. This estimate is three orders of magnitude lower than obtained using the Travis and Arms (1988) model which fails to take into account the importance of metabolism in limiting bioaeeumulation. Consequently, use of such simple models should be avoided to prevent exaggerating phthalate ester exposures via the agricultural foodchain.
SUMMARY 1. A large body of existing data on phthalate ester physieochemieal properties is available. This chemical class exhibits an eight order of magnitude increase in octanol-water partition coefficients (K,~) and four order of magnitude decrease in vapor la'essure (VP) as alkyl chain length increases fi-om I to 13 carbons. Phthalate esters span a twelve order of magnitude change in aqueous solubility over the range of alkyl chain lengths. A critical review of water solubility measurements for higher molecular weight phthalate esters (i.e. alkyl chains ~ 6 carbons) reveals that most published values significantly exceed true values due to exper~ental difficulties associated with solubility determinations for these hydrophobic organic liquids. The higher molecular weight phthaiate esters have extremely low water solubilities.
2. Laboratory and field studies show that partitioning to suspended solids, soils, sediments and aerosols increases as K ~ increases and VP decreases.
723 3. Atmospheric fate processes including photooxidation, washout, and vapor-aerosol partitioning were reviewed. Photodegradation via free radical attack is expected to be the dominant degradation pathway in the atmosphere with predicted half-lives of ca. 1 day for most of the phthalate esters investigated.
4. Biodegradation is expected to be the dominant degradation removal mechanism for phthalate esters entering sewage treatment plants, or released to surface waters, soils and sediments. Numerous studies indicate that phthalate esters are degraded by a wide range of bacteria and actinomycetes under both aerobic and anaerobic conditions. Standardized biodegradation tests with sewage sludge inocula showed phthalate esters undergo ~ 50% ultimate degradation within 28 days with many studies showing ~ 90% removal. Primary degradation half-lives in surface and marine waters range from <1 day to 2 weeks and in soils from <1 week to several months. Longer half-lives may occur in anaerobic, nutrient poor, or cold environments.
5. Numerous experiments have shown that the bioaccumulationof phthalate esters in the aquatic and terrestrial foodchain is limited by biotransformation, which progressively increases with trophic level. In general, bioconcentration factors decrease for organisms that possess more advanced metabolic capabilities. Overall, BCFs span only about three orders of magnitude across phthalate esters. These data indicate that current bioaccumulationmodels developed for persistent organic chemicals that ignore biotransformation grossly exaggerate the bioaccumulation potential of phthalate esters.
6. The results of this study provide the technical basis for future efforts aimed at quantifying multimedia environmental exposure to phthalate esters.
724
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