Wet Peroxide Oxidation of Sediments Contaminated with PCBs

Engineering, University of Wisconsin Madison, Madison, Wisconsin 53706 .... Reaction 6 occurs in the presence of transition metal species (e.g., F...
0 downloads 0 Views 115KB Size

Environ. Sci. Technol. 2000, 34, 3199-3204

Wet Peroxide Oxidation of Sediments Contaminated with PCBs J A M E S E . D U F F Y , * ,†,‡ MARC A. ANDERSON,† CHARLES G. HILL, JR.,§ AND WALTER A. ZELTNER† Water Chemistry Program and Department of Chemical Engineering, University of WisconsinsMadison, Madison, Wisconsin 53706

Wet peroxide oxidation is investigated as a method of treating Hudson River sediments contaminated with polychlorinated biphenyls (PCBs). Aqueous slurries containing 2.5% or 10% (w/w) sediment were oxidized with oxygen and hydrogen peroxide in a 1-L, high-pressure, semibatch reactor at temperatures up to 275 °C. Effluent concentrations of PCBs adsorbed on the sediment and dissolved in the water and gas phases were determined by high-resolution gas chromatography. The rates of oxidation of PCBs were highly dependent on pH, temperature, and the rate of addition of hydrogen peroxide. The addition of hydrogen peroxide significantly increased the rates of oxidation of PCBs over those for conventional wet air oxidation. At 225 °C and a pH of 2.6, addition of hydrogen peroxide at a mass ratio of hydrogen peroxide to sediment of 3:10 resulted in greater than 99% removal of the PCBs as compared to 73% removal for conventional wet air oxidation. The destruction of PCBs increases with decreasing pH over the pH range from 7.5 to 2.6. The optimum temperature for oxidation of PCBs using wet peroxide oxidation was ca. 225 °C. Above this temperature, destruction efficiencies decreased appreciably.

Introduction The presence of sediments contaminated by polychlorinated biphenyls (PCBs) is a major problem in the United States and worldwide. Transport of PCBs from sediments into aquatic food webs sustains contaminant levels in aquatic organisms and provides a route for human exposure. Efficient, cost-effective technologies are needed for remediating these sediments. Although incineration is a proven and widely used technology for treating soils and sediments contaminated with PCBs, there is widespread public opposition to this approach (1). Moreover, incineration of sediments with a high moisture content can be very costly. Other demonstrated technologies used for remediating sediments and soils contaminated with PCBs include thermal desorption, chemical dehalogenation, solvent extraction, and vitrification (1). Each of these alternatives has disadvantages that limit its effectiveness when compared to incineration. Incineration, therefore, remains the remediation technology most often employed at Superfund sites contaminated with PCBs. * Corresponding author phone: (406)994-5926; fax: (406)994-5308; e-mail: [email protected] † Water Chemistry Program. ‡ Present address: Department of Chemical Engineering, Montana State University, Bozeman, MT 59717. § Department of Chemical Engineering. 10.1021/es000924m CCC: $19.00 Published on Web 06/23/2000

 2000 American Chemical Society

Landfilling of contaminated soils and sediments has also been employed at several sites. Public opposition to landfilling, however, is also widespread. Conventional wet air oxidation (WAO) is a commercial process used to remediate aqueous waste streams containing organic solutes (2). WAO utilizes air or oxygen in conjunction with elevated temperatures (150-325 °C) and pressures (3003000 psig) to oxidize highly contaminated waste streams. For a chemical oxygen demand of 15 g/L or more, the reactor operates autothermally. Reactor space times range from 15 min to 2 h depending on the composition of the waste and the extent to which contamination levels are to be reduced. WAO can effectively oxidize most organic contaminants to CO2, H2O, and low molecular weight carboxylic acids (mainly acetic acid) which are resistant to further oxidation. However, halogenated aromatic compounds that do not contain other functional groups (e.g., PCBs and chlorinated benzenes) are resistant to WAO. In a bench scale treatability study conducted by U.S. Filter-Zimpro under guidance from the U.S. Environmental Protection Agency, WAO was evaluated as a potential method for remediating contaminated sediment from the Grand Calumet River/Indiana Harbor Canal (3). Commercial WAO systems are capable of treating solid slurries with solids loadings in excess of 10% (w/w). In this study, a 10% (w/w) slurry of the sediment in water was oxidized with air at 280 °C and 1500 psi for 90 min. The total concentrations of adsorbed polycyclic aromatic hydrocarbons (PAHs) and PCBs were reduced by 99% and 29%, respectively. From these data, EPA personnel concluded that WAO effectively destroys PAHs, but, as anticipated, the WAO process does not destroy PCBs. The oxidation of organic compounds by wet air oxidation is believed to proceed via a free radical mechanism initiated by thermal reactions of the organic substrate with oxygen (4). Free radicals are formed by reactions of oxygen with the weakest C-H or O-H bonds of the organic compounds present. Following this initial step, several reactions occur which involve organic and inorganic radical species and oxygen.

RH + O2 f R• + HO2•


R• + O2 f ROO•


RH + HO2• f R• + H2O2


RH + ROO• f R• + ROOH


Hydrogen peroxide and organic hydroperoxides generated in the above reactions decompose by the following reactions to produce hydroxyl radicals.

H2O2 + (C) f 2HO• + (C)


H2O2 + Mn+ f HO• + OH- + M+(n+1)


ROOH + (C) f RO• + HO• +(C)


In reactions 5 and 7, the collision partner (C) can be either a molecule of water (homogeneous) or a surface of the reactor or sediment (both heterogeneous) (5). The relative importance of the heterogeneous reactions as compared to their homogeneous counterparts depends on the ratio of the surface area of the solids to the volume occupied by the fluid phase, the chemical and physical nature of the solid surfaces, VOL. 34, NO. 15, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY



FIGURE 1. Schematic diagram of the reactor system. and the associated values of the reaction rate constants. Reaction 6 occurs in the presence of transition metal species (e.g., FeII, CuI, CrII, and CoII). An example is the reaction of ferrous iron with hydrogen peroxide, more commonly known as the Fenton’s reagent reaction. The metal species may be dissolved in solution or may be present in the mineral structure of the sediment. As an example of the latter situation, Watts et al. reported that goethite, a common ironcontaining soil mineral, was capable of promoting Fentonlike reactions (6). Fenton-like reactions occur readily at ambient temperature, but reactions 5 and 7 proceed to a substantial degree only at elevated temperatures. The hydroxyl radicals generated in reactions 5-7 are highly reactive. Measured rate constants for the oxidation of most pollutants, including PCBs, by hydroxyl radicals are close to those corresponding to the diffusion-limited reaction (7). Consequently, many researchers have investigated the use of Fenton’s reagent for remediation of aqueous streams, sediments, and soils contaminated with organic compounds (6, 8-12). However, adsorption on particle surfaces has been shown to greatly influence the rate of oxidation of hydrophobic organic compounds by hydroxyl radicals (6, 9). For oxidation of PCBs by Fenton’s reagent in aqueous solutions containing diatomaceous earth, Sedlak and Andren reported that PCBs adsorbed on particulate matter were not appreciably degraded (9). These investigators developed a kinetic model for PCB oxidation that incorporated both partitioning of PCBs between phases and the rates of hydroxyl radical reactions. On the basis of the results of the aforementioned studies, we hypothesized that addition of hydrogen peroxide during wet air oxidation would result in increased production of hydroxyl radicals. Moreover, use of the elevated temperatures characteristic of the wet air oxidation process would increase PCB desorption rates. The combined effect of these processes would lead to increased rates of oxidation of the PCB contaminants. In this study, we have investigated the use of wet peroxide oxidation to treat Hudson River sediment that is contaminated with PCBs. Specific attention is paid to the effects of temperature, pH, and the rate of addition of hydrogen peroxide on the rate and extent of degradation of PCBs adsorbed on the sediment.

Experimental Methods Reactor System. The reactor employed in these studies is shown in Figure 1. It consists of a 1-L high-pressure reactor (Parr Scientific) with connections for pressurization with oxygen, pressure measurement, overpressure protection, depressurization, addition of chemicals, and temperature control. The reactor was placed on a hotplate/stirrer and stirring was accomplished using a 1-in. glass-coated magnetic stirring bar. The surface temperature of the hotplate was controlled and maintained slightly above the reactor tem3200



FIGURE 2. Temperature vs time profile for wet air oxidation and wet peroxide oxidation experiments. perature. The reactor temperature was controlled using a band-heater powered with a temperature controller. The controller was provided with feedback from a thermocouple immersed in the reacting slurry. Hydrogen peroxide or water was added through a dip tube using a high-pressure metering pump to provide accurate addition rates. Overpressure protection was provided by a rupture disk set to relieve at 2000 psi. All reactor components in contact with the slurry were constructed of stainless steel. Preparation of the Slurry. The slurry of sediment was prepared by adding the appropriate amount of dry sediment (National Institute of Standards and Technology: Standard Reference Material 1939a - PCBs in River Sediment) to 400 mL of high purity water (Barnstead Nanopure) to achieve an initial solids concentration of 2.5% (w/w). SRM 1939a is certified reference material sediment originally obtained from the Hudson River. It contains 11.4% (w/w) organic matter and is contaminated with PCBs (approximately 75 mg/kg) and to a much lesser extent with PAHs and pesticides. Sulfuric acid (1 N) or sodium carbonate was added to achieve the desired initial pH. Unless otherwise noted, the initial pH for all experiments was 2.6. Experimental Protocol. After addition of the aqueous slurry, the reactor was sealed and pressurized with oxygen to 300 psi. The stirring rate was set at the desired level (usually 720 rpm), and the slurry was allowed to equilibrate with oxygen for 10 min. After the equilibration period, the oxygen supply valve was closed, and system pressure was monitored for 10 min to ensure that no leaks were present. In the absence of detectable leaks (operationally defined as a noticeable drop in pressure), the reactor and its contents were heated to the temperature of interest. Temperature versus time plots for experiments conducted at various temperatures are shown in Figure 2. Addition of an aqueous solution of hydrogen peroxide at a total volumetric flow rate of 2 mL/min was initiated at 35 min into the experiment and terminated at 85 min. The rate of addition of hydrogen peroxide was varied by adjusting the concentration of hydrogen peroxide in the solution being added. For conventional WAO experiments, high purity water was employed as a control for the experiments with hydrogen peroxide. The overall duration (including heat-up, reaction, and cool-down) for the experiments was ca. 3 h. After cooling to room temperature, the headspace of the reactor was either depressurized through a high-pressure column containing Florisil or sparged through acetone to trap any vapor phase PCBs. The final pH of the slurry was measured prior to further preparation of samples. Sample Preparation. The slurry was filtered through a 0.7 µm glass fiber filter under a slight vacuum into a 1-L

separatory funnel. The sediment and filter [approximately 30% (w/w) water] were transferred to a 250 mL flask, and a recovery spike consisting of PCB congeners 14, 65, and 166 was added. The recovery spike was utilized to monitor for method-related errors but was not used to correct the analytical results. Anhydrous sodium sulfate was then mixed with the sediment. The sediment was extracted with a 1:1 solution of acetone and hexane using a sonication-bath extraction method. The slurry was sonicated for 2 h, allowed to stand overnight to permit batch extraction to occur, and sonicated for a final hour the next morning. The solvent slurry was then filtered through a 0.7 µm glass fiber filter under slight vacuum, and the filtrate was concentrated to ca. 3 mL by rotary evaporator. Finely divided copper was added to the extract during the concentration step to remove elemental sulfur. A 50 mL sample of the aqueous filtrate was transferred to bottles for total organic carbon and metals analyses. A recovery spike was added to the remaining filtrate in the separatory funnel, and the filtrate was extracted using three 50 mL portions of dichloromethane. The extracts were combined and concentrated to ca. 3 mL of hexane using a rotary evaporator. The extracts of the sediment and filtrate were then loaded on cleanup columns (10 mm i.d.) containing 1 in. of sodium sulfate over 3 g of 3.3% deactivated silica gel. The PCBs were eluted with hexane (80 mL) per EPA Method SW-846/3630C. The eluant was concentrated by rotary evaporator or diluted with hexane to an appropriate final volume for analysis. Analytical Procedures. Analyses for specific congeners of PCBs were conducted for all samples using capillary column gas chromatography with electron capture detection (Hewlett-Packard 5890, DB-5 column, and Chemstation software). An internal standard consisting of PCB congeners 30 and 204 was used for quantifying all samples. Both standard congeners were used as retention-time reference points, and congener 30 was used to quantify the sample PCBs. The internal standard congeners are not present in commercial Aroclor mixtures. Total PCB concentrations were estimated by summing individual PCB concentrations obtained using a standard derived from a mixture of Aroclor solutions (13). Calibration using this standard method has been used extensively in the Green Bay Mass Balance study (14). It consists of a 25:18:18 blend of Aroclors 1232, 1248, and 1262 that has been characterized in terms of the weight percentages of all congeners. Selected individual PCB congeners were more accurately quantified using a calibration standard made from solutions of individual pure congeners (Ultra Scientific). Each aqueous filtrate sample was also analyzed for nonpurgeable organic carbon (Shimadzu TOC 5000), and selected samples were analyzed for metals (University of Wisconsin Soil and Plant Analysis Laboratory).

Results and Discussion Validation of Experimental and Analytical Methods. The experimental protocols for extraction of the sediment were validated by comparing average recoveries from duplicate extractions to concentrations of PCBs certified by NIST. Two 10 g portions of sediment were wetted with water [30% (w/ w)], mixed with anhydrous sodium sulfate, and processed according to the procedure described above. Concentrations determined by our protocol may be compared to the certified values for several individual congeners (see Table 1). The agreement between experimental and certified concentrations is quite good, given that some differences would be expected on the basis of differences between extraction methods and analytical procedures. As a check on extraction efficiency, selected sediment residues were extracted a second time using either Soxhlet

TABLE 1. Method Recovery vs Certified Concentration concentration (ng/g) PCB congener no.

certified value

analytical value

recovery %age

52 44 66 118 128 180

4320 1131 840 423 91.2 140.3

3839 1104 960 427 95.3 163

89 98 114 101 104 116

TABLE 2. Recoveries for Control Experimentsa percent recovery percent recovery PCB PCB congener no. control I control II congener no. control I control II 6 19 52 101

99.5 99.0 98.3 98.7

104 100 96.0 90.1

128 180 total

97.5 102 98.8

102 102 98.0

a Control I conditions: temperature ) 25 °C; partial pressure of oxygen ) 300 psi; pH ) 2.6; stirring rate ) 720 rpm. Control II conditions: temperature ) 125 °C; partial pressure of oxygen ) 0 psi; pH ) 2.6; stirring rate ) 720 rpm.

or pressurized fluid extraction techniques. In each case, the second extraction contained less than 10% of the total mass of each PCB congener extracted using sonication, thus demonstrating that the sonication extraction procedure effectively removes the analytes present in each sample. Control Experiments. For wet air oxidation to occur, both elevated temperature and an oxidant (usually oxygen) are required. Two control experiments were performed to verify that PCB losses were not occurring through avenues other than by oxidation. The first control experiment was performed at ambient temperature in the presence of oxygen. The second control experiment was performed at 125 °C in the presence of helium instead of oxygen. All other experimental and analytical conditions and procedures were the same as for the conventional WAO experiments. Table 2 shows recoveries for the two control experiments as a percentage of the untreated sediment extraction values. Total PCB recoveries for the two control experiments were greater than 98%, and individual congener recoveries were generally near 100% with only a few less than 90%. These results indicate that losses of PCBs by routes other than oxidation were negligible. Effect of Stirring Rate on Rates of Conventional WAO. Studies by other research groups indicate that in WAO systems the rate at which oxygen can be supplied may limit the rate of oxidation (15, 16). For batch reactor operation, the rate at which oxygen can be supplied to the slurry is a function of the reactor stirring speed. The effect of oxygen mass transport processes on the overall reaction rate in our system was assessed by measuring the destruction efficiency for PCB-contaminated sediments at several stirring rates. Because magnetic decoupling between the stirring bar and plate was observed at rates in excess of 900 rpm, the influence of stirring rate on the rate of the WAO process was evaluated for rates between 500 and 900 rpm. PCB contaminated sediment [2.5% (w/w) in water] was oxidized at 200 °C and three stirring rates (500, 720, and 870 rpm). Figure 3 shows the percentages of total PCBs and selected individual congeners present in the effluent relative to their influent values. The individual congeners were chosen so as to provide one representative from each of the homologue groups in the range from two to seven chlorine atoms. Inspection of this figure indicates that stirring rate does not appreciably affect PCB degradation over this range. All further studies VOL. 34, NO. 15, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY



FIGURE 3. Wet air oxidation of sediment slurries at stirring rates of 500, 720, and 870 rpm. Experimental conditions: temperature ) 200 °C; sediment loading ) 2.5%; pH ) 2.6.

FIGURE 4. Wet peroxide oxidation of sediment slurries for several rates of addition of hydrogen peroxide. Experimental conditions: temperature ) 200 °C; sediment loading ) 2.5%; pH ) 2.6. reported in this paper were performed at a stirring rate of 720 rpm. Effect of the Addition Rate of Hydrogen Peroxide on the Rates of Destruction of PCBs. Several experiments were conducted to determine a rate of addition of hydrogen peroxide that was effective in degrading the PCBs adsorbed on the sediment. Each of these experiments was performed with a sediment loading of 2.5% (w/w) and an initial pH of 2.6. The rate of addition of hydrogen peroxide was varied by adjusting the concentration in the feed from 0 to 30% (w/w). The percentages of individual congeners and total PCBs remaining after oxidation at 200 °C are presented in Figure 4. On the x-axis, the total mass of hydrogen peroxide added over the duration of the experiment is reported relative to the initial mass of sediment employed. Inspection of Figure 4 indicates that the addition of hydrogen peroxide markedly increases the rates at which all of the congeners are degraded. At 200 °C, the addition of hydrogen peroxide at a mass ratio of peroxide to sediment of 3:10 improved the overall destruction efficiency for PCBs from 76.3% to 99.5%. The concentrations of organic carbon in the effluent filtrates varied in a manner similar to those for the effluent PCBs. For the three experiments displayed in order in Figure 4, final concentrations of organic carbon were 258, 131, and 49 ppmC, respectively. At 250 °C with addition of hydrogen peroxide at a mass ratio of 3:1 (data not shown), nearly complete destruction of all organic matter associated with the sediment and dissolved in the water was observed. Removal of PCBs 3202



FIGURE 5. Wet peroxide oxidation of sediment slurries at several initial pH values. Experimental conditions: temperature ) 200 °C, sediment loading ) 2.5%; mass ratio of hydrogen peroxide to sediment ) 3:10. was 99.9% or greater for all congeners, and the final concentration of organic carbon in the filtrate was 0 ppm-C. Unless otherwise noted, all subsequent experiments reported in this paper were conducted with a mass ratio of hydrogen peroxide to sediment of 3:10. This ratio provided a reasonable compromise between the efficiency of destruction of PCBs and the concomitant costs of using high concentrations of H2O2 in a commercial scale treatment system. The measurable concentrations of the PCB congeners detected in the effluent using this ratio also permits one to assess the effects of various operating parameters on the efficacy of the destruction process. Effect of pH on Rates of PCB Destruction. For both WAO and oxidation with Fenton’s reagent, an optimal pH between 2.5 and 3.5 has often been reported (8, 12, 17). Three wet peroxide oxidation experiments were conducted at 200 °C to determine the effect of pH on the rates of destruction of PCBs. Initial pH values of 2.7, 4.3, and 7.5 were investigated. After oxidation, the corresponding pH values were 3.3, 4.1, and 5.0. Inspection of Figure 5 reveals that the destruction rates for the PCBs of interest increased as the pH decreased. The concentrations of organic carbon in the effluent filtrates followed the same trend. For the three experiments displayed from left to right in Figure 5, the corresponding final concentrations of organic carbon were 49, 96, and 134 ppmC. It is likely that this trend results, at least in part, from a difference in reactivity between the hydroperoxyl and superoxide radicals. These species form an acid/base couple with a pKa of 4.8. The hydroperoxyl radical is the more powerful oxidizing agent. This radical is more reactive than its conjugate base for the oxidation of many organic compounds of biological significance (18, 19). Even though the rates at which hydroperoxyl radicals oxidize PCBs may not be significant, the increased rate of oxidation of generic organic species by hydroperoxyl radicals will result in an increased rate of production of hydroxyl radicals as well as lower concentrations of those compounds which compete with PCBs for reaction with hydroxyl radicals. In addition, at neutral pH, the bicarbonate anion may reduce oxidation rates by scavenging hydroxyl radicals.

HCO3- + HO• f H2O + CO3•-


Compared to the hydroxyl radical, the carbonate radical is much less reactive with organic solutes (20). Effect of Temperature on the Rates of Destruction of PCBs. Temperature is an extremely important variable in

FIGURE 6. Wet peroxide oxidation of sediment slurries at several temperatures. Experimental conditions: sediment loading ) 2.5%; pH ) 2.6; mass ratio of hydrogen peroxide to sediment ) 3:10.

FIGURE 7. Triplicate wet peroxide oxidation experiments. Experimental conditions: temperature ) 225 °C; sediment loading ) 2.5%; pH ) 2.6; mass ratio of hydrogen peroxide to sediment ) 3:10.

the WAO process. System pressure, reaction rates, solubilities of organic compounds and oxygen in water, and rates of desorption of organic compounds from particulate matter are all appreciably affected by temperature. For wet peroxide oxidation experiments, the percentages of total PCBs and selected individual congeners remaining after reaction at temperatures ranging from 125 to 275 °C are depicted in Figure 6. Inspection of these data reveals that the rates of oxidation of the congeners of interest increase with increasing temperature for temperatures up to an optimal temperature of 225 °C, but that above 225 °C, the rates decrease significantly. In fact, wet peroxide oxidation at 275 °C (87% destruction) was only slightly more effective than wet air oxidation at 250 °C (74% destruction). Concentrations of total organic carbon in the filtrate followed a similar trend. Final concentrations of total organic carbon for the five experiments displayed from left to right in Figure 6 were 116, 45, 42, 79, and 140 ppm-C, respectively. Several factors are believed to contribute to the observed trends. As the reaction temperature increases, the contribution of conventional wet air oxidation to the degradation of both PCBs and organic matter becomes more important. The increased rate of wet air oxidation of generic organic matter with increasing temperature results in an increased rate of production of radical species as well as the removal of compounds that compete with PCBs for reaction with hydroxyl radicals. As a result, at high temperatures, addition of hydrogen peroxide becomes more efficient with respect to destruction of PCBs. At lower temperatures, the overall rates of oxidation of PCB congeners may be limited by the rates of desorption of these species from the sediment, especially for the more highly chlorinated congeners. At 125 °C, the large disparities which can be observed in the relative rates of removal of those congeners containing few chlorine atoms compared to their more highly chlorinated counterparts is evidence for the existence of limitations on the rate which are imposed by desorption processes. At temperatures above 225 °C, the rate of decomposition of hydrogen peroxide may exceed that for mixing of different fluid elements within the reactor. The apparent first-order rate constant for reaction 5 varies with both temperature and the ratio of the solid surface area to the liquid volume in the reactor (5).

homogeneous rate constant for decomposition of hydrogen peroxide is 5.1 min-1 (5). The overall rate constant for decomposition of hydrogen peroxide can be much greater than this value under conditions in which accelerated decomposition on reactor and sediment surfaces is prevalent. Under such circumstances, decomposition of hydrogen peroxide may be occurring at a rate that greatly exceeds that of mixing different fluid elements in the reactor. Concentration gradients will exist in the slurry for both hydrogen peroxide and hydroxyl radicals. In regions containing high concentrations of hydrogen peroxide and hydroxyl radicals, significant scavenging of the hydroxyl radicals by hydrogen peroxide can then occur (4, 12).

k ) kh + ks(S/V)


In this equation, kh is the homogeneous rate constant, and ks(S/V) is the surface mediated rate constant. At 275 °C, the

HO• + H2O2 f H2O + HO2•


The hydroperoxyl radical is much less reactive than the hydroxyl radical (18). Moreover, the hydroperoxyl radical may subsequently react with another hydroxyl radical producing water and oxygen. Consequently, the presence of concentration gradients for the hydroxyl radicals and the concomitant increased scavenging of hydroxyl radicals by hydrogen peroxide and hydroperoxyl radicals may contribute to the reduction in the efficiency of PCB degradation observed at very high temperatures. Variability in Experimental Results. To test for variability in experimental results, triplicate experiments were performed at 225 °C, a pH of 2.6, and a mass ratio of hydrogen peroxide to sediment of 3:10. Average values for the percentages of selected individual congeners and total PCBs remaining in the final product mixture are depicted in Figure 7. The error bars represent one standard deviation. For the three experiments, the average percentage of total PCBs remaining was 0.51% with a standard deviation of 0.20%. The fairly large variability for the more heavily chlorinated congeners can be attributed, in part, to the very small concentrations of these congeners in the effluent and the associated uncertainties in the analytical results. Wet Peroxide Oxidation of a 10% (w/w) Sediment Slurry. One experiment was performed with a sediment loading of 10% (w/w) to verify that efficient degradation would occur during the oxidation of slurries containing a sediment loading greater than 2.5% (w/w). This experiment was performed at 225 °C, a pH of 2.6, and a mass ratio of hydrogen peroxide to sediment of 1:4. The total amount of PCBs in the effluent, reported as a percentage of the amount initially charged to the reactor, is comparable to that for the 2.5% (w/w) experiments performed under similar conditions. Overall, 0.67% of the total PCBs remained adsorbed on the sediment, VOL. 34, NO. 15, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY



and less than 0.01% was detected in the water and gas phases. The pH of the final slurry was 3.5, and the concentration of organic carbon in the filtrate was 185 ppm-C. As a check on the extraction efficiency for the higher sediment loading, the sediment residue was extracted a second time using Soxhlet extraction (EPA Method SW-846/3540C). The amount of PCBs extracted using the Soxhlet procedure was less than 10% of that extracted using the sonication bath procedure. The preceding results demonstrate that wet peroxide oxidation may be an effective treatment method for sediments contaminated with PCBs. Fairly simple modifications of commercially available wet air oxidation systems would allow for the treatment of dredged sediments with minimal initial processing of the sediment. Further research directed toward determining the mechanisms, rate-limiting steps, and optimal conditions for wet peroxide oxidation of PCBs is warranted.

Acknowledgments This work was funded by the University of Wisconsin Sea Grant Institute under grants from the National Sea Grant College Program, National Oceanic and Atmospheric Administration, U.S. Department of Commerce, and from the State of Wisconsin. Federal grant number NA46RG0481, project number R/MW-56.

Literature Cited (1) Davila, B.; Whitford, K. W.; Saylor E. S. EPA Engineering Issue; EPA 540-S93-506; U.S. Environmental Protection Agency, 1993. (2) Jackman, A. P.; Powell, R. L. Hazardous Waste Treatment Technologies; Noyes Publications: New Jersey, 1991; pp 90134. (3) U.S. Environmental Protection Agency. EPA 905-R94-007; 1994.




(4) Li, L.; Chen, P.; Gloyna E. F. AIChE J. 1991, 37, 1687-1697. (5) Croiset, E.; Rice, S. F.; Hanush R. G. AIChE J. 1997, 43, 23432352. (6) Watts, R. J.; Udell, M. D.; Monsen, R. M. Wat. Environ. Res. 1993, 65, 839-844. (7) Haag, W. R.; Yao, C. C. D. Environ. Sci. Technol. 1992, 26, 10051013. (8) Sedlak, D. L.; Andren, A. W. Environ. Sci. Technol. 1991, 25, 1419-1427. (9) Sedlak, D. L.; Andren, A. W. Water Res. 1994, 28, 1207-1215. (10) Arnold, S. M.; Hickey, W. J.; Harris, R. F. Environ. Sci. Technol. 1995, 29, 2083-2089. (11) Miller, C. M.; Valentine, R. L.; Roehl, M. E.; Alvarez, P. J. J. Water Res. 1996, 30, 2579-2586. (12) Pignatello, J. J. Environ. Sci. Technol. 1992, 26, 944-951. (13) Mullin M. D. PCB Workshop; U.S. EPA Large Lakes Research Station: Grosse Ile, MI, 1985. (14) Swackhamer D. L. Quality assurance plan - Green Bay mass balance study I. PCBs and Dieldrin; U.S. EPA Great Lakes National Program Office: 1988. (15) Lin, S. H.; Ho, S. J.; Wu, C. L. Ind. Eng. Chem. Res. 1996, 35, 307-314. (16) Sonnen, D. M.; Reiner, R. S.; Atalla, R. H.; Weinstock I. A. Ind. Eng. Chem. Res. 1997, 36, 4134-4142. (17) Chowdhury, A. K.; Ross, L. W. AIChE Symp. Ser. 1975, 71, 4658. (18) Bielski, B. H. J.; Cabelli, D. E. In Active Oxygen in Chemistry; Blackie Academic and Professional: New York, 1995. (19) Bielski, B. H. J.; Cabelli, D. E.; Arudi, R. L.; Ross, A. B. J. Phys. Chem. Ref. Data 1985, 14, 1041-1051. (20) Burns, R. A.; Crittenden, J. C.; Hand, D. W.; Selzer, V. H.; Sutter, L. L.; Salman, S. R. J. Environ. Eng. 1999, 77-85.

Received for review January 21, 2000. Revised manuscript received April 17, 2000. Accepted April 25, 2000. ES000924M