Environ. Sci. Technol. 2000, 34, 4554-4559
Determination of Henry’s Law Constants for Organosilicones in Acutal and Simulated Wastewater M I C H A E L D . D A V I D , * ,†,‡ NICHOLAS J. FENDINGER,§ AND VINCENT C. HAND† Institute of Environmental Sciences, Miami University, 102 Boyd Hall, Oxford, Ohio 45056, The Procter and Gamble Company, Sharon Woods Technical Center, 11511 Reed Hartman Highway, Cincinnati, Ohio 45241, and BASF Agro Research, P.O. Box 400, Princeton, New Jersey 08543
The organosilicone compounds hexamethyldisiloxane (HMDS) and decamethylcyclopentasiloxane (D5) are used in consumer products that are disposed of down the drain and may be released to municipal wastewater treatment facilities. In such systems, their fate is determined largely by partitioning between air and water, which is affected by components of the aqueous phase. The Henry’s law constant (H) is a commonly used expression for a compound’s air-water partitioning behavior. The Equilibrium Partitioning in Closed Systems (EPICS) method of determination of H (dimensionless) was used to assess the fate of HMDS and D5 in wastewater systems. Three chlorinated hydrocarbons and toluene were used to validate the method. The pure water H is measured and reported for D5 and estimated for HMDS, which proved too volatile to be accurately determined using the EPICS technique. The apparent Henry’s law constants (Ha) for all compounds were measured in raw wastewater and simulated wastewater solutions of KCl, humic acid, and resuspended secondary treatment sludge solids. A significant difference between pure water H and Ha in actual and simulated wastewater was observed for all compounds tested. In the simulated wastewater, relationships were observed for organosilicone air-water partitioning over ranges of suspended solids and humic acid and were best correlated to liquid-phase organic carbon content. An equilibrium partitioning model of the system is presented and used to estimate an organic carbon partitioning coefficient (Koc) and H for the organosilicones from the relationship between Ha and liquid-phase organic carbon content. Values of Koc were estimated at 24 000 for D5 and 29 000 for HMDS, with corresponding H values of 3.11 and 32.2 for D5 and HMDS, respectively. Observations from these experiments indicated that the components of natural wastewater can affect the volatility of the test chemicals by a significant amount.
Introduction Volatile polydimethylated siloxanes (VMS) are compounds with a significant vapor pressure (approximately 100 Pa at * Corresponding author present address: BASF Agro Research, P.O. Box 400, Princeton, NJ 08543; phone: (609)716-3200; fax: (609)275-5200; e-mail:
[email protected]. † Miami University. ‡ BASF Agro Research. § The Procter and Gamble Company. 4554
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25 °C), a low water solubility (µg/L range), and a high affinity for solids (log Kd ) 2.2-5.0). VMS fluids consist of -(CH3)2SiO- structural units in either linear or cyclic configurations. Approximately 3% of the total VMS production or about 650 t is used in applications that are disposed of down the drain (1). Disposal pathways from their use in other applications are primarily volatilization to the atmosphere (2, 3). The atmospheric fate of volatile silicones has been studied (4, 5), and they have been classified as nonreactive (6). Although their ultimate fate is volatilization and oxidation in the atmosphere (7-9), partitioning in dynamic wastewater systems between liquid-phase solids, water, and air dictates their short-term fate, which determines their removal pathway from such systems. Previous fate studies for organosilicones have shown them to be volatile and sorptive in aqueous environments (7-10). Studies of removal mechanisms during municipal wastewater treatment (2, 3) attribute approximately equal contributions to removal by sludge adsorption and volatilization. Some fate-determining physical constants for the organosilicones have been reported, but accurate quantifications have been difficult due to their high volatility and hydrophobicity (10). For example, the Henry’s law constant (H) of octamethylcyclotetrasiloxane (D4), for example, has been estimated to be in the range of 3-17 (10, 11). Other important parameters, such as adsorption coefficients, have not been reported for these compounds. The Henry’s law constant is commonly used to quantify a compound’s tendency for volatilization from aqueous systems (12). Determination methods for H include calculation from the ratio of vapor pressure to water solubility and static and dynamic experimental techniques. Direct calculation is limited to those compounds for which accurate physical constants have been determined. The static technique of equilibrium partitioning in closed systems (EPICS), which calculates a dimensionless H (13, 14), is well suited to laboratory-scale environmental fate studies because of the ease of modification of the liquid phase to include salts, solids, organic carbon, and other constituents that affect air-water partitioning. Dynamic determination methods, including the purge-bottle technique (15) and the wetted-wall column (16), are better suited to determination of the pure water H where rapid air-water equilibration can occur. When a fate study requires equilibration between the three phases of an airwater-suspended solid system, the extended equilibration times allowed by the EPICS technique are preferable. The primary disadvantage of the EPICS technique is the high degree of variability with increasing H (17, 18). The fate of organic chemicals in wastewater is affected by volatilization, sorption, and biodegradation. Volatilization, particularly for lower molecular weight chlorinated compounds, can account for up to 100% of the removal pathway from activated sludge facilities (19, 20). Constituents of natural wastewater that can affect air-water partitioning include ionic strength, suspended solids, and humic acid. The “salting out” effect of high ionic strength aqueous solutions has been demonstrated to increase partitioning to air for organics in simulated seawater (21, 22). At lower levels of ionic strength, including ranges seen in wastewater, this effect is less apparent (14, 17). Resuspended activated sludge solids have been demonstrated to bind significant quantities of organophosphate contaminants of wastewater (23). Other wastewater solids have been shown to bind chloroform and methylene chloride (24). Studies of the effect of humic acid at natural ranges on partitioning have been less conclusive, with mixed results being reported (17, 25, 26), including one study that reported an increase in air-water partitioning for 10.1021/es991204m CCC: $19.00
2000 American Chemical Society Published on Web 09/22/2000
organosilicones at humic acid levels of 100 mg/L (10). Higher levels of humic acid in water have been demonstrated to increase aqueous concentration for trichloroethylene (TCE), chloroform, and toluene (27, 28). The purpose of this study is to apply the EPICS method for H determination to the VMS liquids hexamethyldisiloxane (HMDS) and decamethylcyclopentasiloxane (D5) in pure water and wastewater in order to better predict their behavior and fate in wastewater treatment facilities. In addition, the same methodology was used to determine H for HMDS and D5 in simulated wastewater where the composition was varied to determine which factors affect air-water partitioning and thus which aqueous parameters may be important indicators of environmental fate.
Experimental Section Chemicals and Materials. The test chemicals used to validate the methods included methylene chloride, chloroform, TCE, and toluene (Fisher Scientific). The silicone fluids HMDS and D5 were technical grade (Dow Corning). Components added to aqueous systems in order to simulate wastewater included humic acid (Aldrich) and potassium chloride (KCl) (Fisher). Suspended solids used in the wastewater simulation experiment were extracted from composite collections of activated sludge from the aeration basins of three Cincinnati area sewage treatment plants. The three collection sites receive mostly residential wastewater, although each treat some industrial wastes. The sludge liquid was removed in 1.0-L samples into separatory funnels and allowed to settle for 20 min. The settled solids were drawn off into 150-mL centrifuge tubes and centrifuged for 5 min at 3000 rpm. The pellet was then washed by resuspending in pure water and centrifuging again for a total of three washes. The final supernatant wash was disposed, and the solids were freezedried for future resuspension. Before their re-suspension, the solids were heated to 80 °C for at least 2 h to de-activate the bacteria in the solids and prevent biodegradation of test chemicals during analysis. Raw wastewater was influent from the Oxford wastewater treatment plant in Oxford, OH, which receives an average of 2.3 million gal/day of mostly residential flow. EPICS Methodology. In the EPICS method, the ratio of concentrations in the gas phases of two identical headspace sampling vials with different air and water volumes but the same amount of test chemical is used to calculate H. Equation 1 shows the relationship between this concentration ratio and H:
H)
Vw2 - Vw1(R) Vg1(R) - Vg2
(1)
where Vw1&2 are the volumes of water in vials 1 and 2; Vg1&2 are the volumes of headspace in vials 1 and 2; and R is the ratio of headspace concentration of test chemical in vial 1/vial 2. In all H determinations, vials 1 and 2 were prepared in 120-mL headspace sampling vials sealed with gas-tight valves (Supelco). Vial 1 contained 20 mL of liquid (100 mL of headspace); vial 2 contained 80 mL of liquid (40 mL of headspace). After aliquoting the proper volumes of the liquid matrix into each pair of vials, 2.0 µL of methanol with the test chemicals was injected into each vial. The vials were equilibrated for 18-20 h at room temperature (23 ( 1 °C) or by vigorous agitation with a wristaction shaker for 1 h followed by at least 30 min of static equilibration. After equilibration, a 1.0-mL gas sample was removed with a gas-tight syringe from the headspace of each vial and analyzed directly by GC. The concentration ratio, R,
TABLE 1. Data Points Collected in Bivariate Analyses [KCl], M humic acid, mg/L
TSS, mg/L
0.0005
2 40 75 110 150
100 500 1000 1500 2500
X X
0.001
0.005
X
X X X X X
X
0.01
0.02 X
X
X X
was calculated directly from the ratio of chromatographic peak areas, which were assumed to be directly proportional to mass, with no quantitative calibration curves needed. True H was measured in distilled, deionized water for the validation chemicals methylene chloride, chloroform, TCE, and toluene and the organosilicone D5. Each EPICS determination consisted of triplicate pairs of vials. The mean of the headspace concentrations of three 20 mL liquid vials was ratioed to the mean of the headspace concentrations of three 80-mL liquid vials to calculate R for eq 1. Ten of these averaged triplicate determinations were done for the chlorinated chemicals and toluene to validate the method. Five sets of triplicate determinations were then done for D5. Application of the EPICS method to HMDS resulted in high variability. The standard method was therefore modified to estimate H of HMDS. Briefly, the pure water liquid phase was replaced with varying percentages of 2-propanol in water (from 25 to 50%) in order to increase the liquid-phase solubility of HMDS. Two to three EPICS determinations of H were made in each water/2-propanol mixture. The mean calculated H at each point was regressed over percent 2-propanol and extrapolated to the Y axis intercept, which represented pure water and was taken to be the estimated H of HMDS. These modifications are described in detail elsewhere (18). The standard EPICS methodology was applied to actual and simulated wastewater liquid matrixes. In all cases where H was determined in a liquid other than pure water, the resulting parameter calculated is referred to as the apparent Henry’s law constant (Ha). Five sets of triplicate determinations of Ha were done for each of the test chemicals in the Oxford raw wastewater. Effects of conductivity ([KCl]), humic acid, and total suspended solids (TSS) were done experimentally in two separate bivariate analyses. One observed the effect of [KCl] and humic acid on Ha and the other varied [KCl] and TSS. Five levels of each of the variables were chosen to reflect the typical ranges observed in natural wastewater. The 13 points collected in each of these bivariate analyses are described in Table 1. Each data point collected represents triplicate pairs of EPICS vials as was done for the pure water H determinations. Data Analysis A model describing partitioning between the vapor, liquid, and liquid-phase organic carbon was developed. This model is similar to ones described previously for sorption to a solid phase (25, 27). The phase equilibria and their equilibrium constants are described in eqs 2 and 3: Koc
H
Xg {\} Xd {\} Xoc
(2)
Ha
Xg {\} Xl
(3)
where Xg is the test chemical in the gas phase; Xd is the test chemical dissolved in the liquid; Xoc is the test chemical bound to liquid-phase organic carbon; and Xl is the total chemical in the liquid phase (dissolved and bound). VOL. 34, NO. 21, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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When sorption to a liquid-phase component such as organic carbon is a factor, Henry’s law constant determination methods yield Ha. The Ha describes the extent of partitioning between the vapor phase and total liquid phase, which includes both the dissolved chemical in the liquid and the chemical that is bound to organic carbon suspended within the liquid. The corresponding equilibrium constants are given in eqs 4-6:
[X]g
H)
[X]d
)
Mg/Vg Md/Vl
[X]oc
Moc/OC Koc ) ) Md/Vl [X]d Ha )
[X]g [X]l
)
Mg/Vg (Md + Moc)/Vl
(4)
(5)
(6)
where [X]g is the concentration of chemical in gas phase; [X]d is the concentration of chemical dissolved in liquid phase; [X]oc is the concentration of chemical bound to organic carbon; [X]l is the total concentration of chemical in liquid phase (dissoved and bound); Mg is the mass of chemical in gas phase; Vg is the volume of gas; Md is the mass of chemical dissolved in liquid phase; Vl is the volume of liquid; Moc is the mass of chemical bound to liquid-phase organic carbon; and OC is the mass of organic carbon in the liquid phase. Rearranging eq 5 and substituting liquid organic carbon concentration (LOC) ) OC/Vl yields eq 7:
Moc ) Koc × Md × LOC
(7)
Substituting Moc into eq 6 yields eq 8:
Ha )
[X]g Md + (Koc × Md × LOC) Vl
(8)
given Md/Vl ) [X]d:
Ha )
[X]g [X]d(Koc × LOC + 1)
(9)
Given eq 4, the relationship between LOC, Koc, and H can be expressed as
Ha )
H Koc × LOC + 1
(10)
Equation 10 can be rearranged into a linear form:
Koc × LOC 1 1 + ) Ha H H
(11)
Equation 11 shows how the relationship between a liquidphase component which binds test chemical, such as LOC, and the observed partitioning constant Ha can be made linear with a plot of the reciprocal of Ha over the range of LOC with a slope of Kd/H and an intercept of 1/H. In applications to liquids containing humic acid or TSS, a Kd to these liquid components can be calculated by replacing LOC with TSS or humic acid concentration and Koc with Kd in the above model. Thus, the parameters H and Koc or Kd can be estimated from the variation of Ha with the respective liquid-phase component. Statistical Analysis. Standard statistical methods were used for comparison of two means (e.g., comparison of Ha in wastewater to H) and of one data point to a data set (e.g., 4556
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TABLE 2. Henry’s Law Constants from Literature and Determined in Pure Water, Wastewater, Experimental Wastewater Simulations, and Calculateda Ha in simulated wastewater (SD) test chemical lit. Hb
EPICS H (SD)
CH2Cl2
0.089 0.105 (0.013) CHCl3 0.151 0.166 (0.015) TCE 0.396 0.406 (0.021) toluene 0.261 0.263 (0.015) HMDS >19 530 (0.3)e D5 >19 5.46 (1.02)
Ha in humic, TSS, wastewater 2-150 100-2500 calcd c (SD) mg/L mg/L Hd 0.096 (0.02) 0.164 (0.013) 0.364* (0.018) 0.221* (0.026) 64* (57) 0.781* (0.33)
0.100 (0.012) 0.154 (0.011) 0.366* (0.021) 0.250 (0.016)
0.081* (0.008) 0.147* (0.010) 0.358* (0.020) 0.233* (0.015)
0.085 0.105 0.142 0.115 na na
a An asterisk (*) indicates significantly different from pure water H at 95% CI. na, not available. b H of chlorinated chemicals reported by Gossett (5), toluene by MacKay (13), and silicones by Dow-Corning.c N ) 9 for all except HMDS, where N ) 5. d H calculated (vapor pressure/ solubility). e Standard error of regression line reported rather than SD.
comparison of literature values to EPICS-determined H in this study). All references to significant differences between two values refers to determinations made at the 95% confidence level (p < 0.05). Linear regressions, including all reported R 2 values and the slope, intercept, and standard errors for these values, were accomplished using the linest function of Microsoft Excel. Instrumentation. A Hewlett-Packard 5890 series II GC equipped with 30 m × 0.53 mm phenylmethyl polysiloxane column from Restek (Catalog No. 10902) was used for all analyses. Helium carrier gas was used with a flow of 20 mL/ min. All runs were isothermal, but temperature varied with compound mixtures tested. Detection was accomplished with a flame ionization detector. Peaks were electronically integrated for total area. All conductivity measurements were done with a Fisher model 09-325-360 conductivity meter as described in Standard Method 2510B (29). Organic carbon was measured by a Rosemount DC-190 high-temperature total organic carbon analyzer.
Results and Discussion Validation of the EPICS method was accomplished by application to a group of test chemicals with previously measured H. The H of chloroform, methylene chloride, and trichloroethylene had been previously determined using EPICS (14); toluene had been determined using a purgebottle technique (15). Literature values are reported in Table 2 along with the experimental values determined in this study. The determined H for these validation chemicals agree statistically and are within 10% of the reported literature values. Calculated H for the validation chemicals based on measured vapor pressure and water solubility are also reported in Table 2. The calculated H values are generally less than the measured values (up to a factor of 3) but are in the same relative order as they were observed experimentally. The EPICS-determined H value for D5 and HMDS are also reported in Table 2. Application of the EPICS method to D5 resulted in high variability. The relative standard deviation for determination of H of D5 was approximately twice that of the validation chemicals. High variability in repeat determinations of H for very water-insoluble chemicals is an inherent problem with the EPICS method and can cause the H of these compounds to be calculated with extreme
imprecision or even as a negative value (17, 18). This was the case when EPICS was applied to HMDS in pure water and justified the use of the modified EPICS method employed in this study. The plot of Ha over ranges of percent 2-propanol in water was linear (R 2 ) 0.985) and the Y intercept ) 530 is reported in Table 2 as the estimated pure water H for HMDS. The Ha values determined for all test chemicals in raw wastewater are also reported in Table 2. There was no difference in H and Ha in wastewater for chloroform and methylene chloride, but Ha of TCE and toluene were significantly reduced relative to H, by 10 and 16%, respectively. The Ha for both organosilicones in wastewater was more significantly reduced, by 86% in the case of D5. Although HMDS proved to be too hydrophobic to determine H in pure water, partitioning to liquid-phase components of wastewater increased its apparent solubility enough so that Ha could be measured directly with the standard EPICS method. The magnitude of the reduction from H to Ha in wastewater is approximately a factor of 10. A more accurate estimate of the reduction cannot be made because of the high variability in the wastewater Ha determinations for HMDS. All significant differences between H and Ha in wastewater were observed to be reductions, indicating that changes in partitioning were a result of favoring the liquid phase, which is attributable to binding onto solids or liquid-phase organic carbon. We conclude from this observation that for these compounds sorption predominates over ionic strength in determining fate. If ionic strength was predominant, Ha > H would have been observed. To better understand the relationship between wastewater composition and Ha, two experiments were conducted, one varying [KCl] and TSS and the other varying [KCl] and humic acid (Table 1). Increasing [KCl] is predicted to increase Ha linearly (14), as aqueous solubility decreases with increased ionic strength. Equation 11 predicts that 1/Ha will increase linearly with increasing concentrations of TSS or humic acid. Because two different linear relationships were expected for the two variables in each experiment, the data sets were separated into two data groups: one where Ha was determined at five levels of [KCl] while TSS or humic acid concentrations are held constant, and the second group of Ha at five levels of TSS or humic acid while [KCl] was held constant. From each separate data set, linear regressions were performed for Ha against [KCl] and 1/Ha versus TSS or humic acid for each test chemical. There was no significant relationships between [KCl] and Ha for any of the validation chemicals or the organosilicones. The results of the 12 regressions (six chemicals with two [KCl] experiments each) revealed no slope significantly different from zero, and R 2 values ranging from 0.01 to 0.20. Other experiments demonstrating significant partitioning changes with elevated ionic strength have used significantly higher concentrations, up to 30% salt for simulated seawater (21, 22). The results obtained here are consistent with previously reported changes of Ha as a function of ionic strength for organochlorine chemicals, where a change in Ha of less than 1% was observed over the limited ionic strength ranges observed in this study (14). Variations in 1/Ha for the four validation chemicals over ranges of humic acid and TSS were small and not predictable. The slopes of the regressions were not significantly different from zero, and the R 2 values were in the range of 0.01-0.25. Because no observable trends for these chemicals over the ranges of variables were apparent, the mean of all 13 data points from each experiment was determined and reported in Table 2. The two mean values, one for humic acid-[KCl] solutions and one for TSS-[KCl] solutions were compared to the pure water H for each validation chemical. In humic acid-[KCl] solutions, only the mean value for TCE was
FIGURE 1. Ha of D5 over ranges of humic acid and TSS: (9) pure water H determination; (b) Ha determined in humic acid solutions; (O) Ha determined in TSS solutions.
FIGURE 2. Ha of HMDS over ranges of humic acid and TSS: (9) pure water H estimation; (b) Ha determined in humic acid solutions; (O) Ha determined in TSS solutions. significantly different from the pure water mean, being reduced by 10%. In TSS-[KCl] solutions, the Ha for all four validation chemicals was significantly reduced, from 10 to 23%, as compared to H. The Ha values for these chemicals in simulated wastewater are consistent with those observed in actual raw wastewater, with significant differences in partitioning being observed as reductions in the range of 10-20%. We conclude from these observations that components of natural wastewater can affect the fate of these chemicals by decreasing their tendency to volatilize from the liquid phase by a small but significant amount. Variation of Ha over ranges of humic acid and TSS was more significant for the organosilicones than for the organochlorine validation chemicals. Determinations of Ha in pure water, humic acid, and TSS solutions are shown for D5 and HMDS in Figures1 and 2, respectively. The X axes of these figures are aligned to normalize for equivalent organic VOL. 34, NO. 21, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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FIGURE 4. 1/Ha of D5 over ranges of liquid-phase organic carbon. FIGURE 3. Sensitivity analysis: effect of Koc on relationship of H over experimental range of organic carbon. carbon content of the liquid phases from the two experiments. The liquid-phase organic carbon content is calculated from the average measured organic carbon fraction (wt/wt) of 45% for humic acid and 5.5% for TSS. Error bars, representing one standard deviation, are shown in Figures 1 and 2 for all points for which three or more data points were collected. For both of the VMS compounds, relative error of the determinations increases as humic acid and TSS levels are reduced and the liquid-phase approaches pure water. This error is the result of two factors: (i) the inherent variability of the EPICS method at higher H determinations discussed previously and (ii) the difficulty of determining Ha at low organic carbon levels for highly sorptive compounds. This second factor is demonstrated in Figure 3, which is a sensitivity analysis of eq 11 showing the relationship between organic carbon in the range of our experiments and Ha for a compound with an H of 1 and with Koc values ranging from 10 to 100 000. As Koc increases, reduction in Ha becomes more extreme at lower organic carbon levels. This steep slope, especially apparent when Koc exceeds 10 000, is difficult to define experimentally. To calculate the Koc and H for the organosilicones, the data from the humic acid and TSS experiments were combined by calculating the equivalent organic carbon content for the experimental solutions and plotted as 1/Ha over the range of liquid-phase organic carbon. Figures4 and 5 show these data sets for D5 and HMDS, respectively. The Y intercept and the slope of the regression lines of these figures are used to calculate H and Koc, as described in eq 11. These calculated parameters are reported in Table 3. The H of D5 estimated from the liquid-phase organic carbon extrapolation compares well with that determined in pure water with the EPICS method. The H of HMDS determined from the organic carbon extrapolation is approximately a factor of 10 less than the value determined from the 2-propanol extrapolation. Although precise quantification of H for a chemical as volatile as HMDS is difficult to obtain with the EPICS method, it can be concluded from this study that H for HMDS is significantly greater than that of D5, on the order of 10-100× larger, and that HMDS will more rapidly partition to the atmosphere. The results of these experiments demonstrate that components of natural wastewater can affect partitioning relative to pure water systems for the organochlorine and toluene 4558
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FIGURE 5. 1/Ha of HMDS over ranges of liquid-phase organic carbon.
TABLE 3. Koc and H Calculated from Line in Figures 3 and 4 Using Equation 11 test chemical
line regression (R 2)
Koc
H
HMDS D5
0.784 0.771
29000 24000
32.21 3.11
validation chemicals and the VMS test chemicals. Although the organosilicones are volatile, hydrophobic compounds with high Henry’s law constants, the sorption behavior can significantly alter their fate in aqueous systems containing suspended solids or organic carbon. Predictive fate models using a pure water H to estimate behavior that did not account for their high affinity for solids would overestimate the role of volatilization in their removal from wastewater systems.
Acknowledgments This project was supported by The Procter and Gamble Company’s Environmental Safety Division. Technical support and assistance of the personnel in their Ivorydale facility is greatly appreciated. This paper was developed from a Master’s Thesis from Miami University’s Institute of Environmental Sciences.
Literature Cited (1) Allen, R. B.; Kochs, P.; Chandra, G. In The Handbook of Environmental Chemistry. Vol. 3, Part H, Organosilicone Materials; Chandra, G., Ed.; Springer-Verlag: Berlin, 1997; pp 1-25. (2) Parker, W. J.; Shi, J.; Fendinger, N. J.; Monteith, H. D.; Chandra, G. Environ. Toxicol. Chem. 1999, 18, 172-181. (3) Muller J. A.; DiToro, D. M.; Maiello J. A. Environ. Toxicol. Chem. 1995, 14, 1657-1666. (4) Atkinson, R. Environ. Sci. Technol. 1991 25, 863-866. (5) Sommerlade, R.; Pariar, H.; Wrobel, D.; Kochs, P. Environ. Sci. Technol. 1993, 27, 2435-2440. (6) U.S. Environmental Protection Agency. Fed. Regist. 1994, CFR51-59-92. (7) Carpenter, J. C.; Cella, J. A.; Dorn, S. B. Environ. Sci. Technol. 1995, 29, 864-868. (8) Fendinger, N. J.; Mcavoy, D. C.; Eckhoff, W. S.; Price, B. P. Environ. Sci. Technol. 1997, 31, 1555-1563. (9) Xu, S.; Lehmann, R. G.; Miller, J. R. Chandra, G. Environ. Sci. Technol. 1998, 32, 1199-1206. (10) Hamelink, J. L.; Simon, P. B.; Silberhorn, E. M. Environ. Sci. Technol. 1996, 30, 1946-1952. (11) Hobson, J. F.; Atkinson, R.; Carter, W. P. L. In The Handbook of Environmental Chemistry. Vol. 3, Part H, Organosilicone Materials; Chandra, G., Ed.; Springer-Verlag: Berlin, 1997; pp 137-179. (12) Schwartzenbach, R. P.; Geschwend, P. M.; Imboden, D. M. Environmental Organic Chemistry; John Wiley and Sons: New York, 1993. (13) Lincoff, A. H.; Gossett, J. M. In Gas Transfer at Water Surfaces; Brutasaert, W., Jirka, G. H., Eds.; Reidel: Dordrecht, Holland, 1984; pp 17-25. (14) Gossett, J. M. Environ. Sci. Technol. 1987, 21, 202-208. (15) Mackay, D.; Shiu, W.Y.; Sutherland. R. P. Environ. Sci. Technol. 1979, 13, 333-337. (16) Fendinger. N. J.; Glotfelty, D. E. Environ. Toxicol. Chem. 1988, 22, 1289-1293.
(17) Yurteri, C.; Ryan, D. F.; Callow, J. J.; Gurol, M. D. J. Water Pollut. Control Fed. 1987, 59 (11), 950-956. (18) David, M. D. Determination of Henry’s law constants for chlorinated hydrocarbons, toluene, and organosilicones in wastewater matricies. M.S. Thesis, Miami University, Oxford, OH, 1993. (19) Parker, W. J.; Thompson, D. J.; Bell, J. P.; Melcer, H. Water Environ. Res. 1993, 65, 58-65. (20) Bhattacharya, S. K.; et al. Removal and Fate of RCRA and CERCLA Toxic Organic Pollutants in Wastewater Treatment; EPA-600/ 2-89/026; U.S. EPA: Washington, DC, 1989; p 147. (21) Rice, C. P.; Chernyak, S. M.; McConnell, L. L. J. Agric. Food Chem. 1997, 45, 2291-2298. (22) Kucklick, J. R.; Hinckley, D. A.; Bidelman, T. F. Mar. Chem. 1991, 34, 197-209. (23) Villarosa, L.; McCormick, J.; Carpenter, P. D.; Marriott, P. J. Environ. Sci. Technol. 1994, 28, 1916-1920. (24) Dobbs, R. A.; Jelus, M.; Cheng, K. Y. International Conference on Innovative Biological Treatment of Toxic Wastewaters, 1987; pp 585-601. (25) Anderson, M.A. Environ. Sci. Technol. 1992, 26, 2186-2191. (26) Nicholson, B. C.; Maguire, B. P.; Bursill, D. B. Environ. Sci. Technol. 1984, 18, 518-521. (27) Garbarini, D. R.; Lion, L. W. Environ. Sci. Technol. 1985, 19, 1122-1128. (28) Callaway, J. Y.; Gabbbita, K. V.; Vilker, V. L. Environ. Sci. Technol. 1984, 18, 890-893. (29) Standard Methods for the Examination of Water and Wastewater, 17th ed.; American Public Health Association: Washington, DC, 1989.
Received for review October 21, 1999. Revised manuscript received June 20, 2000. Accepted July 19, 2000. ES991204M
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