Degradation of Pentaerythritol Tetranitrate (PETN) by Granular Iron

Despite the somewhat negative effect of the cosolvent on the rate constant, the rate of mass removal can still be enhanced by using a cosolvent. Given...
0 downloads 0 Views 161KB Size
Download PDF
Environ. Sci. Technol. 2008, 42, 4534–4539

Degradation of Pentaerythritol Tetranitrate (PETN) by Granular Iron LI ZHUANG, LAI GUI,* AND ROBERT W. GILLHAM Department of Earth and Environmental Sciences, University of Waterloo, Waterloo, Ontario, Canada N2L 3G1

Received November 28, 2007. Revised manuscript received March 30, 2008. Accepted April 1, 2008.

Pentaerythritol tetranitrate (PETN), a nitrate ester, is used primarily as an explosive. It is of environmental concern, posing a threat to aquatic organisms with an estimated EC50 five times greater than that of RDX. This study evaluated the kinetics and products of PETN degradation in the presence of granular iron. PETN transformation in columns containing 100% granular iron and 30% iron mixed with 70% silica sand followed pseudo first-order kinetics, with average half-lives of 0.26 and 1.58 min, respectively. The reduction pathway was proposed to be sequential denitration, in which PETN was stepwise reduced to pentaerythritol with the formation of pentaerythritol trinitrate and pentaerythritol dinitrate as intermediates. The intermediate of pentaerythritol mononitrate was not detected; however, the nearly 100% nitrogen mass recovery supported complete denitration. Nitrite was released in each denitration step and was subsequently reduced to ammonium by iron. Nitrate was not detected during the experiment, suggesting that hydrolysis was not involved in PETN degradation. Batch experiments showed that when solid-phase PETN is present, dissolution is the rate-limiting factor for PETN mass removal. Using 50% methanol as a cosolvent, PETN solubility was enhanced and thus the removal efficiency was improved. The results demonstrate excellent potential of using iron to remediate PETNcontaminated water.

Introduction Nitrate esters are widely manufactured for two major commercial applications: primarily as explosives, and also as coronary vasodilators (1, 2). 2,2-bis[(nitrooxy)methyl]-1,3propanediol dinitrate (pentaerythritol tetranitrate (PETN), Figure 1) is the most commonly used nitrate ester for these applications. Often, PETN is used in mixture with hexahydro1,3,5-triazine (RDX) to produce Semtex plastic explosive and improvised explosive devices (IEDs) (3) and it has been detected in postexplosion debris (4). Currently, PETN is not vigorously regulated and the threshold limit value (TLV) and maximum workplace concentration (MAK) have not been established. However, a biotoxicity study using the luminescent bacteria test system showed that PETN was five times more toxic than RDX with an EC50 value of 14.54 mg/L/30 min and proposed that it be classified as “toxic to aquatic organisms” (2). The U.S. Department of Defense (DoD) classifies it as a munitions constituent of greatest concern because of its widespread use and potential environmental impact (5, 6). In addition, it was reported that short-term * Corresponding author tel: (519) 888-4567-35207; fax: (519) 7467484; e-mail: [email protected]. 4534

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 42, NO. 12, 2008

exposure may affect the cardiovascular system, resulting in lowering of blood pressure (2, 7). Due to its low vapor pressure (1.035 × 10-10 mmHg) and low Henry’s law constant (1.2 × 10-11 atm · m3 · mol-1) (8), PETN is unlikely to disperse in the ambient air. It is slightly soluble in water (solubility 43 mg · L-1 at 25 °C) (8) and has a low log Kow of 1.61 (8), suggesting weak sorption to organic matter and ease of transport by groundwater. Structurally, the nitrate ester group (C-O-NO2) is analogous to sulfate (C-O-SO3-) and phosphate (C-O-PO32-) ester groups, which are ubiquitous in nature; however, nitrate esters have not been detected as naturally occurring compounds in living organisms (9). They appear to be recalcitrant to natural breakdown and pose a xenobiotic challenge to biological systems (9, 10). Nevertheless, limited studies have shown that PETN was biodegraded in rat urine and feces following sequential denitration processes (11). Binks et al. (9) isolated an Enterobacter cloacae strain PB2 from a soil enrichment under aerobic and nitrogen-limiting conditions and showed that PB2 utilized 2 mol of nitrogen per mol of PETN, producing metabolites of pentaerythritol trinitrate (PETriN) and pentaerythritol dinitrate (PEDN). PEDN was subsequently oxidized to dinitrate dialdehyde. Granular iron is being used increasingly as a cost-effective groundwater remediation alternative because it is capable of reducing a variety of important pollutants (12, 13). The contaminants examined most extensively for treatment by granular iron include chlorinated solvents, azo dyes, nitroaromatic explosives and pesticides, nitrate, and highvalency toxic metals ((14) and references within). The nitro group is a facile electron acceptor and previous studies have shown that iron is capable of reducing organic nitro compounds. Nitrobezene, a nitroaromatic compound, undergoes sequential nitro reduction to aniline by iron under anaerobic conditions, producing intermediates of nitroso and possibly hydroxylamine compounds (15). Both RDX and octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine (HMX) have been shown to degrade to formaldehyde, NH4+, N2O and other soluble products (16–18). Furthermore, glycerol trinitrate (GTN), a compound structurally related to PETN, was shown to undergo sequential denitration with concomitant release of nitrite with 1,2-dinitroglycerin, 1,3-dinitroglycerin, 1-mononitroglycerin, and 2-mononitroglycerin identified as reduction intermediates (14). Even though it is of environmental concern, PETN has received little attention relative to other explosives such as TNT and RDX, and very limited information is available regarding possible methods of remediation. An increased awareness of public health risks and more stringent environmental regulations require the development of effective remediation methods for PETN-contaminated sites. The objective of this study was to evaluate the use of granular iron for degrading PETN in aqueous solution. Using column procedures, we determined the PETN reduction kinetics with iron, identified the reduction intermediates, and proposed the reaction pathways. In batch experiments, when solidphase PETN was present, the effect of dissolution on PETN mass removal was explored and methods for enhancement were considered.

Experimental Section Materials. Granular iron, obtained from Connelly-GPM Inc. (Chicago, IL), was used without pretreatment for the column experiments, but was immersed in either Millipore water or cosolvent solution (50% methanol and 50% water) for 10 days before being used in the batch experiments. The iron 10.1021/es7029703 CCC: $40.75

 2008 American Chemical Society

Published on Web 05/17/2008

FIGURE 1. Structure of pentaerthyritol tetranitrate (PETN). material was characterized previously as containing 89.8% metallic iron (data provided by Connelly-GPM Inc., 1998) with a surface layer of various forms of iron oxides (19). The iron particle size ranged between 0.3 and 2.4 mm with an average specific surface area of 1.02 m2 · g-1, measured by the Brunauer-Emmett-Teller (BET) method using a Micrometrics Gemini III 2375 surface area analyzer. ASTM 2030 sand was purchased from US Silica Company (Ottawa, IL). Before use, the sand was washed with 10% nitric acid, rinsed three times with Millipore water, and oven-dried. Synthesis of PETN. PETN used in the experiments was synthesized in the laboratory by adding pentaerythritol (PE) to an excess nitrating solution consisting of concentrated HNO3 and H2SO4 in a ratio of 2:1 (20). The solution was kept on ice to maintain the temperature below 25-30 °C and was stirred occasionally over a 30 min period. The raw PETN was filtered, washed with DI water, and neutralized with sodium carbonate solution. PETN was purified by dissolving in acetone, recrystallization, and air-drying. The synthesized PETN was verified based on comparison to an analytical standard (0.1 mg · mL-1 in methanol) purchased from AccuStandard Inc. (New Haven, CT), both by retention time on the HPLC chromatogram and by quantification. Column Experiment. The configuration of the two columns used in this study was similar to that described in Gillham and O’Hannesin (12). Each Plexiglas column was 30 cm long by 2.54 cm i.d. with 16 sampling ports distributed along the length. Each sampling port consisted of a nylon Swagelok fitting (0.16 cm o.d.) with a 16G Luer-Lok syringe needle. One column was packed with 100% iron (porosity ) 0.59), and the other was packed with a mixture of 30% iron and 70% silica sand by weight (porosity ) 0.42). The ratios of surface area of iron to solution volume for the 100% and 30% iron columns were 4.70 and 1.37 m2 · mL-1, respectively. PETN solution was prepared by adding crystalline PETN to Millipore water and stirring for 48 h followed by filtration, giving a solution concentration of 6.11 mg · L-1, much below the solubility of 43 mg · L-1 reported in the literature (8). The solution was deoxygenated by sparging with nitrogen gas for approximately 2 h before being delivered to the column. A peristaltic pump was used to pass the solution through the columns at average flow rates of 8.90 and 1.34 mL · min-1 for the 100% and 30% iron columns, respectively. The solution was stored in a carboy with a headspace of nitrogen gas to prevent oxygen invasion. Samples were collected periodically over time and data analysis was performed only after steadystate concentration profiles, i.e., reproducible concentration distributions along the column, had been achieved. Chemical analyses included inorganic ions of NO3-, NO2-, and NH4+, and organic compounds of PETN and potential degradation intermediates. Batch Experiment. Two batch experiments were conducted. The first set (Fe0-H2O) examined the effect of PETN dissolution rate on PETN mass removal by iron. The second set (Fe0-cosolvent) examined the potential for using a cosolvent solution (50% MeOH and 50% H2O) to enhance PETN solubility and thus increase the rate of PETN removal. Before initiating the degradation experiments, each batch vial was prepared in the sequence of weighing the empty vial, adding a specific amount of iron, filling the vial with deoxygenated Millipore water (first set) or cosolvent solution (second set), crimp-sealing with Teflon-lined silicon septum and cap (leaving no headspace), and reweighing. The vials were stored for 10 days to allow autoreduction of most of the

pre-existing passive Fe(III)-oxides on the iron surface to electron-conducting magnetite (19). The pretreatment produces iron with consistent activity and allows comparisons between batch and column experiments. To reduce abrasion of the iron surfaces during mixing, the iron particles were held in place by magnets, attached on opposite sides of the vials. The Fe0-H2O experiment consisted of 4 sets of vials in duplicate, each set with a different Fe/H2O ratio and 8 mg of PETN (concentration of 200 mg · L-1). Iron in amounts of 20, 10, 5, or 1 g was added to 60 mL glass vials, resulting in solution/iron ratios (g · g-1) of approximately 3, 6, 12, and 60. The Fe0-cosolvent experiment consisted of 5 sets of 40 mL glass vials in duplicate, with each pair containing 10 g of iron and 20, 35, 50, 65 or 80 mg · L-1 of PETN. The reduction reaction was initiated by addition of PETN stock solution in acetonitrile (10,000 mg · L-1) via needle injection through the septum into the individual vials. In the Fe0-H2O experiment, 8 mg of PETN was introduced to all vials by spiking with the stock solution and allowing the PETN to recrystallize. In the second batch experiment, a specific amount of PETN stock solution was injected into each set of vials to achieve the desired initial concentrations (20, 35, 50, 65, and 80 mg · L-1). Because of the cosolvent effect, PETN remained in solution. All vials were loaded onto an orbital rotator at 50 rpm. At particular elapsed times, duplicates were sacrificed for organic and inorganic analyses. PETN was analyzed for both batch tests and NO3-, NO2-, and NH4+ were analyzed only for the Fe0-H2O test. Chemical Analyses. Concentrations of nitrate and nitrite were analyzed using a Dionex ion chromatograph (Dionex ICS 2000) equipped with a conductivity detector, an ioneluent generator, and a Dionex AS-40 autosampler. A Dionex IonPac AS18 column (4 × 250 mm) was used. The injection volume was 25 µL and the mobile phase was 30 mM KOH at a flow rate of 1.2 mL · min-1. The detection limit for both nitrite and nitrate was 0.50 mg · L-1. Ammonium (NH4+-N) analysis was performed using the phenate method described in Standard Methods for the Examination of Water and Wastewater (21). The absorbance at a wavelength of 630 nm was measured using a Beckman DU 530 UV/vis spectrophotometer with a light path of 1 cm. The detection limit was 0.10 mg · L-1. Analyses for PETN and intermediate products were performed using a series 1100 Hewlett-Packard highperformance liquid chromatograph (HPLC) equipped with a UV-visible diode array detector, a quaternary pump, and an autosampler. A Zorbax SB-C18 column (3.5 µm particles, 4.6 × 150 mm) and a Zorbax guard column (5 µm particles, 4.6 × 12.5 mm) were used. Aqueous samples were centrifuged in 1.5 mL vials for 5 min at 10,000 rpm prior to being loaded onto the autosampler. A water-methanol-acetonitrile mixture (40:50:10, v/v/v) was used as the mobile phase at a flow rate of 1.0 mL · min-1. The addition of 10% acetonitrile to the eluent sufficiently changed the polarity of the mobile phase such that the separation of PETN and its degradation intermediates was improved and the peaks were sharpened. The injection volume was 100 µL and the absorbance was measured at a wavelength of 210 nm. The detection limit for PETN was 0.10 mg · L-1. A series of PETN external standards was prepared using the analytical standard and analyzed with the samples for PETN quantification. Because standards for the intermediates were not commercially available, the intermediates were not quantified. To identify unknown intermediates, fractions of unknown peaks appearing on the HPLC chromatogram were collected at the corresponding retention times through repeated injections and concentrated by air-drying. The samples were redissolved in acetonitrile and analyzed using the positive ion ammonia chemical ionization method, with a JEOL HX110 VOL. 42, NO. 12, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

4535

FIGURE 2. Changes in PETN concentration in the column using 30% iron and 70% silica sand. double focusing mass spectrometer operated at a mass resolution of ∼1000. The source temperature was 200 °C with an electron energy of 200 eV. All samples were introduced by direct injection and heated when necessary.

Results and Discussion Degradation Kinetics of PETN. The measured PETN concentrations (C) were normalized to the initial concentration (C0) and plotted as C/C0 versus residence time, converted from distances along the column based on the porosity and measured flow rates following the normal procedure (e.g., ref (12)). PETN reduction in the 100% iron column was consistent with pseudo first-order kinetics, with R2 values (correlation coefficient of a least-squares fit) greater than 0.96 for the profiles taken after steady-state conditions had been reached. The average rate constant (with standard deviation, SD) was 2.70 ( 0.22 min-1, giving a half-life of 0.26 ( 0.020 min and a surface area-normalized rate constant (kSA) of (5.73 ( 0.47) × 10-4 L · m-2 · min-1. Due to the extremely fast degradation rate in the 100% iron column, there may be a significant level of uncertainty in the calculated rate constant. Consequently, the amount of iron was reduced to 30% in the second column. Figure 2 shows four representative PETN concentration profiles measured between 380 and 650 pore volumes (PV), where one PV is the total volume of void space (solution) in the column, well after steady-state concentrations had been achieved. The pseudo first-order model fit the data well (R2 ) 0.96-0.97), with an average rate constant ( SD of 0.44 ( 0.029 min-1, giving a half-life of 1.58 ( 0.10 min. The corresponding kSA ( SD was (3.20 ( 0.22) × 10-4 L · m-2 · min-1. The kSA value is similar to that obtained for the 100% column and is in reasonable agreement with kSA values reported in the literature for other explosives that have been tested with similar iron materials. For example, Oh et al. (14, 18) reported values of 2.75 ( 0.50 × 10-4 L · m-2 · min-1, 6.00 ( 1.07 × 10-4 L · m-2 · min-1, and 1.29 ( 0.14 × 10-3 L · m-2 · min-1 for nitroglycerin, RDX, and HMX reduction, respectively. Nitrogen Mass Balance. Concentrations of NH4+, NO2-, and NO3-, in addition to PETN, were measured to establish the nitrogen mass balance during the column experiments. No NO3- was detected over the entire experiment. Figure 3 is typical of the concentration profiles of PETN, NH4+, and NO2- (expressed in relative moles of N) along the column, taken from the 30% iron column after 420 PV. PETN concentration decreased to the detection limit (100 µg · L-1) within 10 min. Concurrently, the concentration of NO2reached a maximum of 49% of the initial N then began to decline. Meanwhile, NH4+ continued to accumulate over time. NO2- gradually decreased to below the detection limit and the amount of NH4+ approached a maximum toward the end of the column. Throughout the column, the nitrogen mass recovery, i.e., the ratio of the sum of the N in nitrogenous 4536

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 42, NO. 12, 2008

FIGURE 3. Masses of nitrogenous compounds (PETN, NO2-, and NH4+) during PETN reduction in the column using 30% iron and 70% silica sand. Profiles were taken after 420 PV. compounds including PETN, NH4+, and NO2- to the initial N in PETN, expressed in percentage, ranged between 84% and 104%. In particular, toward the end of the column, all of the N in the initial PETN was present as NH4+, indicating that all the nitro groups on PETN were reductively removed by the release of nitrite, which was subsequently reduced to NH4+, as the final nitrogen-bearing product. The reduction of NO2- by iron, producing NH4+, has been previously reported (22). A reproducible dip in the nitrogen mass balance occurred at a residence time of 4-5 min (Figure 3). The deficit was attributed to unidentified nitrogen-bearing intermediates. Three unknown peaks were observed on the HPLC chromatograms during PETN analysis, at retention times of 1.3 (peak 1), 2.3 (peak 2), and 3.9 (peak 3) min (cf. 6.5 min for PETN) (Figure S1a). All unknown peaks showed similar trends of an initial increase to a maximum followed by a decrease, with eventual disappearance for peaks 2 and 3 and a plateau for peak 1 before exiting the iron column (Figure S2a and b). The pattern of appearance and disappearance of the unknown peaks suggests that they are intermediate products of PETN degradation. Peaks 3, 2, and 1 reached their maximum values at 4.1, 5.8, and 8.5 min, suggesting a sequential formation of intermediates. We noted that the area of peak 1 was much greater relative to that of peak 2 and 3. This was a consequence of the interference from nitrite. The similar profile of nitrite measurement by IC to peak 1 by HPLC (Figure S2b) indicates the presence of nitrite in the HPLC analysis; on the other hand, the dissimilarity at maximum and before exiting the iron column suggests the coexistence of peak 1 rather than nitrite itself (Figure S2b). Though a longer residence time, allowing for complete nitrite removal, would have been helpful, the data suggests that unknown 1 may have approached a constant value toward the end of the column. It was not possible to determine the relative rates of degradation of PETN and the intermediate products, however, the short duration of the dip in the nitrogen balance (approximately 5 min) indicates that the intermediates degraded quickly. Identification of Intermediate Products. The three unknown peaks, believed to be intermediate products of PETN reduction, were identified as PE (unknown 1), pentaerythritol dinitrate (PEDN, unknown 2), and pentaerythritol trinitrate (PETriN, unknown 3) according to the LC-MS analysis described in Supporting Information. Because analytical standards of PEDN and PETriN were not available and because nitrite coeluted with PE, the identified intermediates could not be quantified in this study. However, the transient nature of the identified PEDN and PETriN explains the reproducible early time deficit in the nitrogen mass

FIGURE 4. Proposed pathway for degradation of PETN by granular iron. balance since the occurrence of their maximum peak areas coincides with the lowest values in the nitrogen mass balance curve. Degradation Pathways. Based on the identified intermediate products, PETriN and PEDN, and the carbon-bearing end product, PE, as well as the near 100% nitrogen mass balance, it is proposed that PETN degradation in the presence of iron follows a sequential reductive degradation pathway with the release of NO2- in each denitration step, as shown in Figure 4. The nitrite liberated from the stepwise reductive reactions was further reduced by iron, as in eq 1. Equation 2 indicates that, stoichiometrically, a total of 32 mol of electrons are required to degrade 1 mol of PETN to NH4+ and PE as the final reduction products. Thus, considering Fe2+ to be the final iron product, 16 mol of Fe0 would be required. NO2- + 6e- + 6H2O f NH4+ + 8OH-

(1)

C5H8(ONO2)4 + 32e- + 28H2O f C5H8(OH)4 + 4NH4+ + 36OH- (2) The apparent absence of pentaerythritol mononitrate (PEMN) in the sequential denitration pathway may be because (i) the rate of conversion of PEMN to PE was faster than its formation from PEDN; (ii) the amount of PEMN was very small, particularly with the low initial PETN concentration, resulting in the concentration being below the detection limit of HPLC analysis; or (iii) the analytical method used in this study was not sensitive to PEMN. Even though the lack of PEMN made the last denitration step in the degradation pathway uncertain, the complete nitrogen recovery indicated that all nitrite groups were removed. Hydrolysis did not appear to be involved since NO3- was not detected. If NO3- were released, its slower reduction rate with iron compared to NO2- would have made it detectable (22). The absence of nitrate as a degradation product is also supported by the absence of nitrate in the later batch experiment conducted at an initial concentration of 200 mg · L-1 PETN. The proposed pathway for PETN degradation is similar to that of glycerol trinitrate, another nitrate ester compound. Oh et al. (14) showed that in the presence of cast iron, glycerol trinitrate was stepwise reduced to 1,2- and 1,3-dinitroglycerins, then to 1- and 2-mononitroglycerins, and finally to glycerol and NH4+. The reduction process was proposed to

FIGURE 5. Results of batch tests with solid-phase PETN (200 mg · L-1) and 20, 10, 5, and 1 g iron: Changes in (a) aqueous PETN concentrations and (b) masses of nitrogenous compounds (PETN, nitrite, and ammonium) over time. be reductive rather than hydrolytic, consistent with the observations of this study. Effect of Dissolution on Mass Depletion. In spite of the high reduction rate constant, if solid-phase PETN is present, rates of mass depletion may be limited by the relatively low solubility (6.11 mg · L-1 was achieved in this study). The first batch experiment described in the methods section, in which the initial PETN concentration (200 mg · L-1) exceeded its solubility, was performed to evaluate this possibility. Figure 5a and b summarize the average concentrations of PETN in the solution phase and the relative mass of nitrogen in the reduction products (sum of N in NO2- and NH4+ divided by N in initial PETN) during PETN reduction for the tests conducted with different loadings of iron: 20, 10, 5, and 1 g. When crystalline PETN is present, the concentration of PETN in solution is a consequence of depletion by degradation and addition by dissolution. The decrease in concentration within the first 3 days (Figure 5a) represents the period when the degradation and dissolution processes were coming to equilibrium. As a common trend for all cases, following the initial period of decline, the aqueous concentrations became moderately steady (below solubility), suggesting equilibrium between the rates of degradation and dissolution. Notably, the “equilibrium” concentration in the solution phase decreased with increasing iron loading. As a consequence of more iron, the degradation rate of PETN was increased, leading to lower aqueous concentrations, higher concentration gradients, and thus higher rates of dissolution. This leads to higher rates of mass depletion, as reflected in the total amount of PETN that was degraded over time (Figure 5b). There is an approximately linear increase in the concentration of nitrogenous degradation products after 2 or 3 d, particularly in the tests with 10, 5, and 1 g of iron, corresponding with the period during which the PETN concentration in solution was relatively constant. The trend was less apparent in the test with 20 g of iron, as a consequence of more rapid and complete reduction of PETN VOL. 42, NO. 12, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

4537

before achieving steady state. Furthermore, the rate of mass depletion (accumulation of degradation products) increased with the iron loading. For example, with 20 g of iron, there was almost complete removal of PETN by day 4, while with 1 g of iron, there was only 44% removal by day 8. Using the same experimental approach and amount of iron (10 g), the first-order surface-area normalized rate constant for degrading aqueous PETN (6.02 mg · L-1) was observed to be 8.81 × 10-3 L · m-2 · h-1, giving rate constants of 4.49, 2.25, 1.12, and 0.23 h-1 for 20, 10, 5, and 1 g iron tests, respectively. Using the first-order degradation model and assuming all the added PETN was in the solution phase, the reduction of 200 mg · L-1 PETN to the detection limit (0.10 mg · L-1) would require 1.69, 3.38, 6.77, and 33.81 h for the tests with 20, 10, 5, and 1 g of iron, respectively, which are far shorter than the observed times for mass depletion. It is clear that if solid-phase PETN is present, the rate of mass depletion is limited by the dissolution process. This limitation could be overcome to some degree by increasing the surface area concentration of iron. Enhancement by Cosolvent. When methanol (50%) was used as a cosolvent to enhance PETN solubility, very similar pseudo first-order disappearance of PETN was observed (Figure S3) in the five batches, which were initiated with different PETN concentrations (20, 35, 50, 65, and 80 mg · L-1). PETN was degraded to below the detection limit within 12 h in all batches at an average half-life ( SD of 1.68 ( 0.38 h. This result provides further evidence that dissolution rate is a limiting factor. The results obtained from the cosolvent experiment show that with the loading of iron used in this study, and within the concentration range from 20 to 80 mg · L-1, the rate of PETN mass removal from solution can be increased when the dissolved PETN concentration increases. Therefore, cosolvent may have the potential for improving the rate of depletion by means of increasing PETN solubility. This could, for example, have application to soil washing treatment as demonstrated in a previous study (23). Cosolvents, though effective in increasing mobility and solubility of hydrophobic contaminants, may also pose negative effects on degradation rates. To explore this possibility, a supplemental batch test using 10 g of iron and aqueous PETN (6.02 mg · L-1), in the presence and absence of cosolvent, was conducted. The kSA was 8.81 × 10-3 and 1.60 × 10-3 L · m-2 · h-1 for PETN reduction in the Fe0-H2O and Fe0-cosolvent systems, respectively, indicating the degradation rate in the Fe0-H2O system to be about 5 times greater than that in the Fe0-cosolvent systems. The difference suggests a decrease in degradation rate caused by the presence of cosolvent. In a similar experiment, Clark et al. (24) found that the dechlorination rate by granular iron decreased with increasing cosolvent fractions. The decrease in degradation rate was attributed to a reduction in contaminant sorption on the iron surface. Despite the somewhat negative effect of the cosolvent on the rate constant, the rate of mass removal can still be enhanced by using a cosolvent. Given the kSA values of 1.60 × 10-3 and 8.81 × 10-3 L · m-2 · h-1 in the Fe0-cosolvent and Fe0-H2O system, 1.60 × 10-3 and 8.81 × 10-3 L of PETNcontaminated water can be completely treated in unit time with unit iron surface area in the Fe0-cosolvent and Fe0-H2O system, respectively. As a consequence, PETN mass removal in the Fe0-cosolvent system would be 0.13 mg (1.60 × 10-3 L of 80 mg · L-1 PETN-contaminated water) whereas the mass depleted in the Fe0-H2O system would be 0.05 mg (8.81 × 10-3 L of 6 mg · L-1 PETN-contaminated water). Therefore, using a cosolvent to increase PETN solubility can be an effective means to improve treatment efficiency when crystalline PETN is present. 4538

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 42, NO. 12, 2008

In summary, the results demonstrate that, in the presence of iron, PETN in the aqueous phase can be reduced rapidly and effectively to pentaerythritol and NH4+ through sequential denitration. When present in crystalline form, rates of mass depletion can be greatly enhanced by introducing a cosolvent. In addition, PETN may be present in a mixture with other munitions such as RDX and TNT. It is possible that these nitro explosives will compete for reactive iron surfaces and thus affect PETN degradation kinetics; however, all nitro explosives can be effectively treated by iron. PETN is synthesized using pentaerythritol and nitric acid with sulfuric acid as a catalyst. Hence, in wastewater facilities, one should expect high concentrations of nitrate and sulfate. Our previous studies showed that the presence of high concentrations of nitrate may result in iron passivation (25, 26). Therefore, when using iron for the treatment of PETN, the effect of nitrate should be taken into consideration.

Acknowledgments Financial support for this research was provided through the NSERC/Dupont/EnviroMetal Industrial Research Chair in Groundwater Remediation held by R.W.G. We thank Dr. Richard Smith in the Department of Chemistry at the University of Waterloo for assistance with the LC/MS analysis.

Supporting Information Available HPLC chromatogram showing PETN and its unknown degradation intermediates, LC-MS spectra of neat PETN and PE as well as unknown intermediates and their identification, formation and subsequent disappearance of PETN degradation intermediates, PETN degradation in cosolvent batch experiment. This information is available free of charge via the Internet at http://pubs.acs.org.

Literature Cited (1) Bucher, J. R. NTP toxicology and carcinogenesis studies of pentaerythritol tetranitrate (CAS No.78-11-5) with 80% D-lactose monohydrate (PETN, NF) in F344/N Rats and B6C3F1 Mice (Feed Studies); National Toxicology Program Technical Report Series Vol. 365; Environmental Health Information Service, 1989; pp 1-192. (2) Drzyzga, O.; Gorontzy, T.; Schmidt, A.; Blotevogel, K. H. Toxicity of explosives and related cmpounds to the luminescent bacterium Vibrio fischeri NRRL-B-11177. Arch. Environ. Contam. Toxicol. 1995, 28, 229–235. (3) Ballesteros-Rueda, L. M.; Herrera-Sandoval, G. M.; Mina, N.; Castro-Rosario, M. E.; Briano, J. G.; Hernandez-Rivera, S. P. Spectroscopic signatures of PETN part II: detection in clay. Proc. SPIE 2006, 6217, 62173D/1-8. (4) Tamiri, T.; Zitrin, S.; Abramovich-Bar, S.; Bamberger, Y.; Sterling, J. GC/MS analysis of PETN and NG in post-explosion residues. In Proceedings of the 4th International Syposium on Advances in Analysis and Detection of Explosives; Yinon, J., Ed.; Kluwer: Dordrecht, Netherlands, 1993; pp 323-334. (5) U.S. Department of Defense Operational Rangers. More reliable clean up cost estimates and a proactive approach to identifying contamination are needed; GAO-04-610, 2004. (6) Wollin, K.-M.; Dieter, H. H. Toxicological guidelines for monocyclic nitro-, amino- and aminonitroaromatics, nitramines, and nitrate esters in drinking water. Arch. Environ. Contam. Toxicol. 2005, 49, 18–26. (7) IPCS, CEC. Pentaerythritol tetranitrate; 2005; available at http:// www.inchem.org/documents/icsc/. (8) Hazardous Substances Data Bank (HSDB); United States National Library of Medicine: Bethesda, MD, 2003; available at http:// toxnet.nlm.nih.gov/cgi-bin/. (9) Binks, P. R.; French, C. E.; Nicklin, S.; Bruce, N. C. Degradation of pentaerythritol tetranitrate by Enterobacter cloacae PB2. Appl. Environ. Microbiol. 1996, 62, 1214–1219. (10) White, G. F.; Snape, J. R. Microbial cleavage of nitrate esters: defusing the environment. J. Gen. Microbiol. 1993, 139, 1947– 1957. (11) King, S. Y. P.; Fung, H. L. Rapid microbial degradation of organic nitrates in rat excreta: re-examination of the urinary and fecal metabolite profiles of pentaerythritol tetranitrate in the rat. Drug Metab. Dispos. 1984, 12, 353–357.

(12) Gillham, R. W.; O’Hannesin, S. F. Enhanced degradation of halogenated aliphatics by zero-valent iron. Ground Water 1994, 32, 958–967. (13) Blowes, D. W.; Ptacek, C. J.; Benner, S. G.; McRae, C. W. T.; Bennett, T. A.; Puls, R. W. Treatment of inorganic contaminants using permeable reactive barriers. J. Contam. Hydrol. 2000, 45, 123–137. (14) Oh, S.-Y.; Cha, D. K.; Kim, B. J.; Chiu, P. C. Reduction of nitroglycerin with elemental iron: pathway, kinetics, and mechanisms. Environ. Sci. Technol. 2004, 38, 3723–3730. (15) Agrawal, A.; Tratnyek, P. G. Reduction of nitro aromatic compounds by zero-valent iron metal. Environ. Sci. Technol. 1996, 30, 153–160. (16) Hundal, L. S.; Singh, J.; Bier, E. L.; Shea, P. J.; Comfort, S. D.; Powers, W. L. Removal of TNT and RDX from water and soil using iron metal. Environ. Pollut. 1997, 97, 55–64. (17) Singh, J.; Comfort, S. D.; Shea, P. J. Remediating RDXcontaminated water and soil using zero-valent iron. J. Environ. Qual. 1998, 27, 1240–1245. (18) Oh, S.-Y.; Cha, D. K.; Kim, B. J.; Chiu, P. C. Reductive transformation of hexahydro-1,3,5-trinitro-1,3,5-triazine, octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine, and methylenedinitramine with elemental iron. Environ. Toxicol. Chem./SETAC 2005, 24, 2812–2819.

(19) Ritter, K.; Odziemkowski, M. S.; Gillham, R. W. An in situ study of the role of surface films on granular iron in the permeable iron wall technology. J. Contam. Hydrol. 2002, 55, 87–111. (20) Sam Barro’s PowerLabs, available at ttp://www.powerlabs.org/ chemlabs/petn.htm. (21) Standard Methods for the Examination of Water and Wastewater, 19th ed.; American Public Health Association: Washington, DC, 1985. (22) Alowitz, M. J.; Scherer, M. M. Kinetics of nitrate, nitrite, and Cr(VI) reduction by iron metal. Environ. Sci. Technol. 2002, 36, 299–306. (23) Jawitz, J. W.; Sillan, R. K.; Annable, M. D.; Rao, P. S. C.; Warner, K. In-situ alcohol flushing of a DNAPL source zone at a dry cleaner site. Environ. Sci. Technol. 2000, 34, 3722–3729. (24) Clark, C. J., II; Rao, P. S. C.; Annable, M. D. Degradation of perchloroethylene in cosolvent solution by zero-valent iron. J. Hazard. Mater. 2003, 96, 65–78. (25) Ritter, K.; Odziemkowski, M. S.; Simpgraga, R.; Gillham, R. W.; Irish, D. E. An in situ study of the effect of nitrate on the reduction of trichloroethylene by granular iron. J. Contam. Hydrol. 2003, 65, 121–136. (26) Lu, Q.; Gui, L.; Gillham, R. W. Effects of nitrate on trichloroethylene degradation by granular iron. Earth Sci. Frontiers 2005, 12, 176–183.

ES7029703

VOL. 42, NO. 12, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

4539