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Published online 7 November 2005
Published in J Environ Qual 34:2208-2216 (2005)
DOI: 10.2134/jeq2005.0032
© 2005 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America
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TECHNICAL REPORTS

Bioremediation and Biodegredation

Biotic and Abiotic Degradation of CL-20 and RDX in Soils

Fiona H. Crockera,*, Karen T. Thompsonb, James E. Szecsodyc and Herbert L. Fredricksonb

a Analytical Services, Inc., 3532 Manor Dr., Suite 3, Vicksburg, MS 39180
b U.S. Army Engineer Res. and Dev. Center, CEERD-EP, 3909 Halls Ferry Rd., Vicksburg, MS 39180
c Pacific Northwest National Lab., P.O. Box 999 K3-61, Richland, WA 99354

* Corresponding author (Fiona.H.Crocker{at}erdc.usace.army.mil)

Received for publication January 26, 2005.

   ABSTRACT
TOP
 ABSTRACT
INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 REFERENCES
 
The caged cyclic nitramine 2,4,6,8,10,12-hexanitro-2,4,6,8,10,12-hexaazaisowurtzitane (CL-20) is a new explosive that has the potential to replace existing military explosives, but little is known about its environmental toxicity, transport, and fate. We quantified and compared the aerobic environmental fate of CL-20 to the widely used cyclic nitramine explosive hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) in surface and subsurface soil microcosms. Soil-free controls and biologically attenuated soil controls were used to separate abiotic processes from biologically mediated processes. Both abiotic and biological processes significantly degraded CL-20 in all soils examined. Apparent abiotic, first-order degradation rates (k) for CL-20 were not significantly different between soil-free controls (0.018 < k < 0.030 d–1) and biologically attenuated soil controls (0.003 < k < 0.277 d–1). The addition of glucose to biologically active soil microcosms significantly increased CL-20 degradation rates (0.068 < k < 1.22 d–1). Extents of mineralization of 14C–CL-20 to 14CO2 in biologically active soil microcosms were 41.1 to 55.7%, indicating that the CL-20 cage was broken, since all carbons are part of the heterocyclic cage. Under aerobic conditions, abiotic degradation rates of RDX were generally slower (0 < k < 0.032 d–1) than abiotic CL-20 degradation rates. In biologically active soil microcosms amended with glucose aerobic RDX degradation rates varied between 0.010 and 0.474 d–1. Biodegradation was a key factor in determining the environmental fate of RDX, while a combination of biotic and abiotic processes was important with CL-20. Our data suggest that CL-20 should be less recalcitrant than RDX in aerobic soils.

Abbreviations: 2ADNT, 2-amino-4,6-dinitrotoluene • CL-20, 2,4,6,8,10,12-hexanitro-2,4,6,8,10,12-hexaazaisowurtzitane • HMX, octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine • HPLC, high pressure liquid chromatography • k, first-order degradation rate constant • PLFA, phospholipid fatty acids • RDX, hexahydro-1,3,5-trinitro-1,3,5-triazine • TNT, 2,4,6-trinitrotoluene


   INTRODUCTION
TOP
 ABSTRACT
 INTRODUCTION
MATERIALS AND METHODS
 RESULTS
 DISCUSSION
 REFERENCES
 
THE RECENTLY DEVELOPED energetic compound 2,4,6,8,10,12-hexanitro-2,4,6,8,10,12-hexaazaisowurtzitane (CL-20; Fig. 1) has a greater specific energy than 2,4,6-trinitrotoluene (TNT), hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX; Fig. 1), or octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine (HMX) (NAWCWD, 1998; Simpson et al., 1997). Since the use of CL-20 in munitions is still in the experimental stage, little is known about the environmental fate and toxicity of CL-20. Military training activities on firing ranges have led to the contamination of soils and ground water with TNT, RDX, and HMX (Clausen et al., 2004; Jenkins et al., 2001; Pennington et al., 2003; Walsh et al., 2001, 2004). Such activities along with manufacturing and storage are likely routes by which CL-20 may contaminate soils and ground water. The CL-20 has been shown to be highly toxic to earthworms (Eisenia andrei) (Robidoux et al., 2004), but nontoxic to soil and marine bacteria, a green alga, or terrestrial plants (Gong et al., 2004). Therefore, the ability to predict the fate and environmental impact of explosives, such as CL-20, in surface and subsurface soils is important to sustainable military range practices.



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  		Fig. 1. Molecular structures of hexanitrohexaazaisowurtzitane (CL-20) and hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX).

 
Recently, several abiotic reaction processes have been shown to transform CL-20. For instance, CL-20 is photoreactive (Hawari et al., 2004; Szecsody et al., 2004), is degraded by Fe0 (Balakrishnan et al., 2004), and it undergoes alkaline hydrolysis more rapidly than RDX or HMX (Balakrishnan et al., 2003; Qasim et al., 2004). In addition, the abiotic reactivity of CL-20 under aerobic conditions varies widely with a number of different sediments, minerals, clays, and Fe and Mn oxides. Abiotic degradation half-lives of CL-20 ranged from minutes to hundreds of hours in contrast with RDX, which was not degraded under the same conditions (Szecsody et al., 2004). Furthermore, soils with low water contents exhibited slower abiotic degradation rates of CL-20 than soils with higher water contents, indicating that water promoted cleavage of CL-20. The CL-20 also appears to be unstable in neutral aqueous solutions in contact with glass, and agitation increases this slow rate of decomposition (Monteil-Rivera et al., 2004; Szecsody et al., 2004).

Abiotic degradation of CL-20 has been shown to produce a variety of different end products and the product ratios formed depended on the specific abiotic process, possibly indicating different pathways. Photolysis of CL-20 produced the end products nitrite and nitrate, ammonia, formic acid, and trace amounts of nitrous oxide, N2 gas, and glyoxal (Hawari et al., 2004). Alkaline hydrolysis and degradation by Fe0 produced nitrite, nitrous oxide, ammonia, and formic acid (Balakrishnan et al., 2003, 2004; Szecsody et al., 2004), while glyoxal and then glycolate were additionally formed with Fe0 (Balakrishnan et al., 2004). Abiotic transformation of CL-20 by sediments, minerals, and clays generated nitrite, nitrate, and formate (Szecsody et al., 2004), while nitrite, nitrous oxide, and formate were formed in aqueous CL-20 solutions (Monteil-Rivera et al., 2004). A general degradation pathway that involves denitration (removal of two nitro groups) followed by ring cleavage of the "roof" or "attic" C–C bond was proposed for most of these abiotic processes. The abiotic degradation mechanism of Okovytyy et al. (2005) differs at this point by proposing the continued release of all nitro groups, via HONO or NO2· elimination, to form a stable aromatic end product, 1,5-dihydrodiimidazo[4,5-b:4'5'-e]pyrazine.

A few studies have shown that CL-20 can be biodegraded in aerobic surface soils (Jenkins et al., 2003; Trott et al., 2003), by bacteria isolated from these soils (Bhushan et al., 2003; Trott et al., 2003), and by functionally diverse bacterial enzymes (Bhushan et al., 2004a, 2004b). Similar to abiotic degradation processes, nitrite, ammonium, nitrous oxide, formate, and glyoxal were among the various end products formed (Bhushan et al., 2003, 2004a, 2004b). The enzymatic studies with purified salicylate 1-monooxygenase (Bhushan et al., 2004b) and nitroreductase (Bhushan et al., 2004a), and a membrane associated flavoenzyme (Bhushan et al., 2003) have shown that the biotransformation of CL-20 is catalyzed by an oxygen-sensitive, one-electron transfer reaction that caused one nitrite ion to be released from the molecule. Two one-electron transfer reactions were proposed to release two nitrite ions and destabilize the ring leading to ring cleavage.

Studies comparing the coupled biological/abiotic transformation of CL-20 and RDX in the same soils appear to be lacking. In the above studies, 14C–CL-20 mineralization was not investigated, due to the unavailability of 14C-labeled CL-20 at the time. Due to the complex and strained molecular structure of CL-20, we hypothesized that CL-20 would be more susceptible to aerobic transformation reactions than RDX. In this study, we examined the biological and abiotic degradation of CL-20 and RDX, and the mineralization of uniformly labeled 14C–CL-20 in aerobic surface and subsurface soil microcosms. Our objectives were to quantify and compare the rates of degradation and thus determine the relative importance of biological vs. abiotic degradation of CL-20 and RDX in these aerobic soils. This data will help to define differences in degradation mechanisms that affect the environmental fate and impact of these two explosives.


   MATERIALS AND METHODS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
RESULTS
 DISCUSSION
 REFERENCES
 
Soils
Surface and subsurface soils were obtained from China Lake, CA (January 2003 and October 2003), and from active firing ranges at Fort Bliss, TX (June 2002), and Fort Polk, LA (July 2003). China Lake is a Naval Air Warfare Center Weapons Division site located about 250 km north of Los Angeles, CA. Soil at China Lake was collected from Site 8, an area that receives surface water runoff and is therefore an active fluvial deposit. This area may have been contaminated with explosive residues in the past; however, we did not detect any explosive residues in this soil via high pressure liquid chromatography (HPLC) analysis (see below). Subsurface soil was collected from China Lake Site 7 at a depth of 4.5 m in January 2003. Fort Bliss is located adjacent to El Paso, TX. Soil was collected from the center of a detonation crater (Soil A) and next to a low-order (incomplete) detonation (Soil B). Explosive residues were not detected in Soil A, whereas Soil B was contaminated with 2,4,6-trinitrotoluene (TNT; 3649 mg kg–1) and 2-amino-4,6-dinitrotoluene (2ADNT; 177 mg kg–1). Fort Polk is located 100 km north of Lake Charles, LA. Soil collected at Fort Polk was from a grenade range and did not contain any detectable explosive residues. Selected physical and chemical characteristics of the surface and subsurface soils used in this study are shown in Table 1.


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  		Table 1. Selected chemical and physical properties of soils.

 
All soils were stored at 4°C and homogenized at the obtained field water content before use. Surface soils were passed through a 2-mm sieve to remove small pebbles and rocks. The China Lake surface soil collected in October 2003 had very low moisture content and so it was amended with sterile water to the same moisture content (6% w/w) as the soil collected in January 2003 and incubated for 2 to 3 d at room temperature to stimulate microbial activity in the soil.

Soil Microcosms
Soil microcosms consisted of 15-mL screw cap test tubes containing 1 g of soil plus 5 mL of a N-free mineral salts medium buffered to pH 7.0. The medium composition was similar to that described by Shelton and Tiedje (1984) and consisted of (per liter): KH2PO4, 0.272 g; K2HPO4, 0.348 g; MgSO4·7H2O, 0.2 mg; FeSO4·7H2O, 2 mg; CaCl2·2H2O, 0.03 mg; MnCl2·4H2O, 0.5 mg; H3BO3, 0.05 mg; ZnCl2, 0.05 mg; CuCl2, 0.03 mg; Na2MoO4·2H2O, 0.01 mg; CoCl2·6H2O, 0.5 mg; NiCl2·6H2O, 0.05 mg; Na2SeO3, 0.5 mg. The NH4Cl, resazurin, NaHCO3, Na2S·9H2O, and vitamins were omitted and the medium was kept aerobic by not sparging the medium and the headspace of microcosms with O2–free N2 gas. Glucose (1 g L–1) was added to the indicated microcosms as an exogenous C source, and CL-20 (3 or 10 mg L–1) or RDX (3 or 10 mg L–1) was added as the N source. The CL-20 and RDX were prepared as methanol or acetone stock solutions, respectively. Medium concentrations of CL-20 or RDX at 3 mg L–1 were obtained by adding a small volume (<0.5 mL L–1) of explosive stock solution to the bottom of a sterile flask and allowing the solvent to evaporate before the addition of the sterile mineral salts medium. The medium was then stirred overnight to allow the CL-20 or RDX to dissolve. Medium concentrations of 10 mg L–1 were obtained by adding 10 µL of the appropriate explosive stock solution (0.2% v/v) to 5 mL of mineral salts medium in each test tube. To quantify abiotic degradation rates associated with soil minerals and clays, biologically attenuated soil control microcosms were used. These microcosms contained sterile mineral salts medium and either soil that had been autoclaved three consecutive times (1 h at 121°C) or soil that was amended with HgCl2 (final concentration, 300 mg L–1). Abiotic soil-free control microcosms, containing only sterile mineral salts medium, were included to quantify reactions caused by the glassware and medium components (Szecsody et al., 2004). Microcosms were continually mixed on a tissue culture rotator (Glas-Col, Terre Haute, IN) at 30 rpm and incubated in the dark at 25 ± 1°C.

Analyses
Two to three replicate microcosms were analyzed at selected time intervals. The total amount of energetic remaining (aqueous plus soil-sorbed) was determined using an organic extraction consisting of the addition of 5 mL of methanol (CL-20) or acetonitrile (RDX) to each microcosm. Microcosms were placed into an ultrasonic bath (Branson, Danbury, CT) for 18 h at 15°C. Following centrifugation at 2800 x g for 10 min, the supernatant was filtered through a 0.45-µm PTFE filter and analyzed by HPLC.

The CL-20 and RDX were analyzed by HPLC using an Agilent 1100 Series HPLC (Palo Alto, CA) equipped with a quaternary pump, autosampler with a 200-µL loop injector, diode array UV absorbance detector, and a column oven. A Hypersil ODS reverse-phase C-18 HPLC column (100 by 4.6 mm; 5-µm particle size) was used as the primary column, along with a Hypersil ODS C-18 guard column (20 by 4 mm; 5-µm particle size). The system was operated at 39°C and at a flow rate of 1.5 mL min–1. The isocratic mobile phase consisted of 68% 20 mM NH4Cl, 31.4% methanol, and 0.6% butanol. The CL-20 was measured at 234 nm and RDX was measured at 254 nm. The calibration curves of CL-20 and RDX were linear between 0.1 and 50 mg L–1 for 1:1 (v/v) solutions of solvent and water.

Carbon-14–CL-20 Mineralization
Carbon-14–labeled CL-20 was used to investigate the rate and extent of CL-20 mineralization in the China Lake surface and subsurface soils. Microcosms were prepared in 125-mL serum bottles containing 5 g of soil, 50 mL of the mineral salts medium (±glucose, 1 g L–1), and an inner vial (10 by 75 mm) containing 1 mL of 1 M KOH. The serum bottles were sealed with Teflon-coated butyl rubber stoppers and aluminum crimp seals. The mineral salts medium was amended with both unlabeled and 14C-labeled CL-20 to achieve an initial concentration of 3.5 mg L–1 and 1.44 µCi L–1. The microcosms were incubated in the dark at 25 ± 1°C with constant shaking at 175 rpm. Control microcosms included soils amended with HgCl2 (300 mg L–1) or only the sterile mineral salts medium. At each sampling time, the bottles were opened and the KOH removed for liquid scintillation counting. In addition, 0.6 mL of the aqueous phase was collected to determine the aqueous concentration of CL-20. The aqueous sample was mixed with an equal volume of acidified acetonitrile (Monteil-Rivera et al., 2004), filtered, and analyzed by HPLC (above). A fresh volume of KOH (1 mL) was added to the inner vial and the microcosm sealed. At the end of the incubation period, the soil was collected by centrifugation (5000 x g for 10 min) and the radioactivity in the aqueous phase (1-mL) and soil were measured. One gram of soil was mixed with 5 mL of acidified acetonitrile and then the soil was treated as described above. The filtered soil extract was diluted with an equal volume of water and then analyzed by HPLC and liquid scintillation counting. The acetonitrile-extracted soil was air-dried and the amount of soil-bound 14C-labeled CL-20 or 14C-labeled CL-20 degradation products remaining was determined by combustion to 14CO2 using a Model 307 Packard Sample Oxidizer (Packard Co., Meriden, CT).

Biomass Determinations
At Day 0 and Day 21, microbial biomass estimates of the China Lake soils were obtained from the concentration of ester-linked phospholipid fatty acids (PLFA). Phospholipid fatty acids analyses were performed according to the procedures of Balkwill et al. (1988) and White and Ringelberg (1998).

Chemicals
The CL-20 was obtained from A.T.K. Thiokol Propulsion (Brigham City, UT) as the {epsilon}-isomer (>99% pure). The [UL-14C]CL-20 was synthesized at Thiokol from glyoxal-14C and was found to have a specific activity of 0.345 mCi mmol–1 as determined by HPLC and liquid scintillation counting. The RDX (>95% pure) was synthesized by Stan Caulder of the Indian Head Division, Naval Surface Warfare Center, Indian Head, MD.

Degradation Kinetics
The concentrations of CL-20 and RDX were plotted as C/Ci vs. time (t), where C is the concentration of energetic remaining at time, t (d), and Ci is the initial concentration of energetic at time 0. To determine k, the first-order degradation rate constant (d–1), the degradation curves were fit according to first-order kinetics as follows:

[1]
where Y is the value of C/Ci at a given time, t, and Ymax is the maximum value of C/Ci at time 0. This general model was adapted to provide better fits to some of the data that exhibited lag phases and/or incomplete degradation as follows:

[2]
where tL is the time when the lag phase ends, Lag is the maximum value of C/Ci during the lag phase, and Ymin is the minimum, measured value of C/Ci. Values for k were calculated for each curve using GraphPad Prism version 4.00 for Windows, GraphPad Software (San Diego, CA). Values for tL, Ymin, Ymax, and Lag were either calculated or were set to initial values to achieve the best curve fit as determined by coefficients of fit values (r2) and how well the curve-fit reflected the actual data. If values for tL and Ymin were set to 0, then Eq. [2] collapsed to Eq. [1] (first-order degradation kinetics without a lag phase and complete degradation). If only the value for tL was set to 0, the equation modeled the situation with no lag phase and incomplete degradation. The degradation half-life (d) was then calculated from the following equation:

[3]
The rate data were analyzed using the general linear models procedure (PROC GLM) of SAS (SAS Institute, 2003).


   RESULTS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
DISCUSSION
 REFERENCES
 
Biodegradation of CL-20 in China Lake Surface and Subsurface Soils
The CL-20 was biodegraded in China Lake surface and subsurface soil microcosms incubated under aerobic conditions (Fig. 2) . In the surface soil, the biodegradation of CL-20 was significantly faster in soil amended with glucose than in soil without glucose (Fig. 2A, Table 2). The CL-20 was completely biodegraded in the glucose-amended soil after 13 d, whereas 60 to 65% of the CL-20 was degraded in the unamended soil after 21 d. Similarly, glucose enhanced the rate and extent of biodegradation in the subsurface soil (Fig. 2B, Table 2). The rates of biodegradation were faster in the surface soil microcosms compared with the subsurface soil (Fig. 2, Table 2). In the glucose-amended soil microcosms, the rate of biodegradation was 2.5 times faster in the surface soil (half-life = 1.7 d) than in the subsurface soil (half-life = 4.2 d). In the unamended soils, there was a twofold difference in CL-20 biodegradation rates between the surface and subsurface China Lake soils (Table 2). Before incubation, microbial biomass estimates of the surface soil (1.76–2.94 x 1010 cells kg–1 [wet wt.]) and subsurface soil (0.99–2.38 x 1010 cells kg–1 [wet wt.]) were similar (Table 3). After 21 d in the presence of CL-20, an increase in biomass was observed in both surface and subsurface soil microcosms with or without glucose, suggesting that CL-20 was being used to support growth of the indigenous soil microorganisms. The microbial biomass was about three times higher in the surface soil microcosms (1.48 x 1011 cells kg–1 [wet wt.]) than the subsurface soil microcosms (5.31 x 1010 cells kg–1 [wet wt.]), probably as a result of the faster biodegradation rate in the surface soil microcosms. The surface soil had slightly higher NO–3–N and clay contents than the subsurface soil (Table 1), which may have contributed to higher microbial activities in the surface soil.



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  		Fig. 2. Degradation of hexanitrohexaazaisowurtzitane (CL-20; 10 mg L–1) in China Lake soil microcosms incubated under aerobic conditions. (A) Surface soil; (B) subsurface soil. Symbols: unamended soil, open circles; soil + glucose, solid triangles; soil + glucose + HgCl2, open triangles; mineral salts medium + glucose, solid squares.

 

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  		Table 2. First-order rate constants (k), half-lives, and correlation coefficients (r2) for hexanitrohexaazaisowurtzitane (CL-20; 10 mg L–1) and hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX; 10 mg L–1) degradation in China Lake soils.

 

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  		Table 3. Microbial biomass estimates in China Lake soil microcosms incubated with hexanitrohexaazaisowurtzitane (CL-20; 10 mg L–1) or hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX; 10 mg L–1).

 
Biodegradation of RDX in China Lake Surface and Subsurface Soils
Under aerobic conditions, RDX was completely biodegraded after 8 or 13 d of incubation in the surface soil amended with or without glucose, respectively (Fig. 3A) . When the surface soil was amended with glucose, the rate of RDX biodegradation (half-life = 1.5 d) was similar to the rate of CL-20 biodegradation under these conditions (Table 2). The RDX biodegradation rate (half-life = 4.7 d) was significantly faster than the CL-20 biodegradation rate (half-life = 17.3 d) when glucose was not added to the surface soil microcosms (Table 2). In the surface soil microcosms amended with RDX, with or without glucose the microbial biomass estimates increased nearly 10-fold from 1.37 to 2.45 x 1010 cells kg–1 (wet wt.) on Day 0 to 1.65 to 1.95 x 1011 cells kg–1 (wet wt.) after 21 d of incubation (Table 3). In the subsurface soil microcosms, RDX was biodegraded slowly with approximately 15% of the RDX transformed in 21 d in soil amended with glucose, whereas there was no apparent biodegradation in the subsurface soil that did not receive glucose (Fig. 3B, Table 2). There were only minor changes in microbial biomass in these microcosms during the incubation period (Table 3).



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  		Fig. 3. Degradation of hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX; 10 mg L–1) in China Lake soil microcosms incubated under aerobic conditions. (A) Surface soil; (B) subsurface soil. Symbols: unamended soil, open circles; soil + glucose, solid triangles; soil + glucose + HgCl2, open triangles; mineral salts medium + glucose, solid squares.

 
Abiotic Degradation of CL-20 and RDX
The CL-20 was prone to abiotic degradation in addition to biological degradation under aerobic conditions (Fig. 2). This abiotic degradation occurred in the China Lake soil microcosms amended with HgCl2 and in microcosms containing only sterile mineral salts medium. The rates of abiotic CL-20 degradation were 0.009 to 0.030 d–1 (half-lives = 23.1–77 d) and were not significantly different between soils amended with HgCl2 and the sterile mineral salts microcosms (Table 2). Furthermore, the abiotic rates of CL-20 degradation were similar to the biological rate of CL-20 degradation in China Lake soil microcosms that did not receive glucose. In contrast, RDX was not degraded by abiotic mechanisms in aerobic surface or subsurface soils treated with HgCl2 or in microcosms only containing sterile mineral salts medium (Fig. 3, Table 2).

Biodegradation of CL-20 and RDX in Various Surface Soils
For a broader perspective on the fate of CL-20 in soils, we compared the degradation of CL-20 in the China Lake surface soil with three firing range soils (Tables 2 and 4). The rates of CL-20 biodegradation in biologically active soil microcosms amended with glucose varied greatly between surface soils from China Lake, Fort Polk, and Fort Bliss (0.068 < k < 1.222 d–1). For each soil type, rates of CL-20 degradation were significantly faster in active soils than in biologically attenuated soils (Tables 2 and 4). The organic C content, pH, NO–3–N, or clay content of the soils (Table 1) could not explain the different CL-20 biodegradation rates in these soils. Similarly, rates of RDX biodegradation were very variable among the four surface soils amended with glucose (0.017 < k < 0.474 d–1) and not related to the rate of CL-20 biodegradation in the same soil (Tables 2 and 4). In active surface soil microcosms amended with glucose, the fastest CL-20 biodegradation rate occurred with the Fort Bliss B soil followed in order by China Lake, Fort Polk, and Fort Bliss A soil. With respect to RDX, the order of biodegradation rates was as follows: China Lake > Fort Bliss A soil > Fort Bliss B soil.


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  		Table 4. First-order rate constants (k) for hexanitrohexaazaisowurtzitane (CL-20) and hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) degradation in various surface soils.

 
Mineralization of Carbon-14–CL-20 in China Lake Soils
The aerobic mineralization of CL-20 in biologically active China Lake surface and subsurface soil microcosms was extensive with between 41.1 and 55.7% of the initial 14C–CL-20 recovered as 14CO2 after 41 d of incubation (Fig. 4 , Table 5). The production of 14CO2 was slower than the biodegradation of CL-20, since 14CO2 did not appear until after a lag phase of between 2 and 7 d (Fig. 4). Lag phases were not evident from the CL-20 disappearance data (Fig. 5) . In the subsurface soil amended with glucose, CL-20 was completely degraded in 4 d, but the maximum extent of 14CO2 evolved was not reached until after 14 d. Similarly, CL-20 was completely biodegraded in the surface soil microcosms treated with or without glucose in 18 d, but the extent of 14CO2 evolution only approached the maximum after 41 d. While biodegradation of CL-20 appeared to be similar between the glucose-amended surface soil and the unamended surface and subsurface soils, the distribution of 14C-label from 14C–CL-20 was different and may simply reflect that mineralization had progressed further toward completion in the surface soil microcosms. In subsurface soil microcosms without glucose, a slightly lower amount of 14CO2 was evolved (41.1% vs. 50.8–55.7%), and larger amounts of the 14C-label were recovered from the aqueous phase (27.1% vs. 9.7–16.4%) and soil (24.2% vs. 10.5–12.1%; Table 5).



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  		Fig. 4. Cumulative mineralization of 14C-hexanitrohexaazaisowurtzitane (14C-CL-20; 3.5 mg L–1, 1.44 µCi L–1) to 14CO2 in China Lake soil microcosms. (A) Surface soil; (B) subsurface soil. Symbols: unamended soil, open circles; soil + glucose, solid triangles; soil + glucose + HgCl2, open triangles; mineral salts medium + glucose, solid squares.

 

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  		Table 5. Carbon (14C) balance of 14C-hexanitrohexaazaisowurtzitane (14C–CL-20; 1.44 µCi L–1) mineralization in China Lake soil microcosms.

 


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  		Fig. 5. Change in CL-20 concentration in China Lake soil microcosms amended with 14C-hexanitrohexaazaisowurtzitane (14C–CL-20; 3.5 mg L–1, 1.44 µCi L–1). (A) Surface soil; (B) subsurface soil. Symbols: unamended soil, open circles; soil + glucose, solid triangles; soil + glucose + HgCl2, open triangles; mineral salts medium + glucose, solid squares.

 
Little mineralization of CL-20 was observed in the HgCl2–amended soil microcosms or sterile mineral salts medium microcosms (Fig. 4). Between 0.46 to 1.3% of the 14C-label was evolved as 14CO2 in the HgCl2–amended soils, whereas 5.8% 14CO2 was produced in the sterile mineral salts medium microcosms (Table 5). In contrast, the abiotic degradation of CL-20 in these control microcosms was extensive (54–90% degraded) after 41 d of incubation (Fig. 5). The higher agitation rate used with these microcosms was probably responsible for the increase in abiotic degradation compared with our earlier results (Fig. 2, 5). The majority of the 14C-label (52.7–79.1%) was found as unidentified, water-soluble products (Table 5).


   DISCUSSION
TOP
 ABSTRACT
 INTRODUCTION
 MATERIALS AND METHODS
 RESULTS
 DISCUSSION
REFERENCES
 
This study has shown that CL-20 is susceptible to both biotic and abiotic degradation in aerobic soils. Biological rates of CL-20 degradation were relatively rapid (half-lives = 0.6–31.5 d) in both surface and subsurface soils (Tables 2 and 4), suggesting that CL-20 will probably not persist in soils. Similar rates of CL-20 biodegradation were observed with garden and agricultural soils (Trott et al., 2003) and with Fort Polk soil that was moist, but unsaturated (Crocker et al., unpublished data, 2005). However, significantly slower rates of CL-20 degradation were observed with three firing range soils incubated under moist, unsaturated conditions (half-lives = 144–686 d; Jenkins et al., 2003). It was not apparent from the limited number of soils and soil properties measured in this study or in Jenkins et al. (2003), which soil property might be responsible for the variable rates of CL-20 biodegradation observed in these studies.

The biodegradation of CL-20 and RDX was stimulated by the addition of a readily available C source, glucose (Fig. 2 and 3, Table 2). Since microorganisms use explosives as N sources for growth, they require an adequate source of C and energy (Binks et al., 1995; Coleman et al., 1998; Seth-Smith et al., 2002). A variety of exogenous C substrates have been shown to stimulate RDX, HMX, or TNT biodegradation in other soils (Fuller et al., 2004; Speitel et al., 2001; Waisner et al., 2002). The soils used in this study had low organic C contents (<1%), and so the metabolic activity of the indigenous microorganisms was probably low. Since biomass estimates were higher in China Lake soil microcosms treated with glucose than without glucose and in which degradation of CL-20 or RDX occurred (Table 3), glucose stimulated the growth and activity of microorganisms with the catabolic potential to transform CL-20 or RDX.

A prior explosive history did not appear to be necessary for degradation of CL-20, since CL-20 was degraded in all of the soils used in this study. In addition, the ability of a soil to biologically degrade CL-20 did not always correlate with the ability to degrade RDX. In the absence of any soils contaminated with CL-20, soils that were potentially contaminated with explosive residues, such as from military training ranges and ordnance manufacturing sites, were used. Generally, degradation of new xenobiotic compounds, like CL-20, requires a period of acclimation, the length of which depends on the contaminant structure and how similar the structure is to existing compounds. Soils with a history of explosive contamination should select for microbial populations adapted to biodegrade existing explosives and thus have the potential to biodegrade CL-20 without a long acclimation period. Since the China Lake surface soil may have been in contact with energetic residues, this contamination may have influenced the more rapid energetic degradation rate observed relative to the uncontaminated subsurface soil. Of the soils collected, only the soil from around a low-order detonation (Fort Bliss B soil) had detectable levels of the explosives, TNT and 2ADNT, but not RDX. The fastest rate of CL-20 biodegradation occurred in this soil, but RDX was very slowly degraded (Table 4). On the other hand, the soil collected at Fort Bliss from the site of a detonation crater with no detectable levels of energetics (Fort Bliss A soil) had a faster rate of RDX biodegradation than the rate of CL-20 biodegradation. This data suggests that different microbial populations apparently degrade these two explosives.

Abiotic degradation of CL-20 in biologically attenuated aerobic soil microcosms and in microcosms with only sterile mineral salts medium was observed at significant rates (0.003 < k < 0.277 d–1). The abiotic rate of CL-20 degradation in the China Lake surface soil microcosm amended with HgCl2 (half-life = 23.1 d) was similar to the abiotic rates (18–32 d half-life) observed by Szecsody et al. (2004) for this soil. We observed little to no significant difference between the abiotic rates of CL-20 degradation in the biologically attenuated soil microcosms or mineral salts medium controls (Fig. 2, Tables 2 and 4). This may have been due to the reactivity of CL-20 with the glass tubes, the neutral mineral salts medium, or agitation of the microcosms with only a minor contribution by the soil to the abiotic degradation rate.

In contrast, RDX was not abiotically degraded in biologically attenuated soil or sterile mineral salts medium microcosms incubated under aerobic conditions (Fig. 3, Tables 2 and 4). The absence of abiotic degradation of RDX has been observed in other soils (Sheremata et al., 2001; Szecsody et al., 2004). The highly strained nature of the CL-20 molecule compared with RDX may explain why CL-20 is more easily transformed under abiotic, aerobic conditions and why CL-20 has less thermal and frictional stability (Monteil-Rivera et al., 2004).

Mineralization of 14C–CL-20 to 14CO2 in the biologically active China Lake surface and subsurface soil microcosms was evidence of ring cleavage and the subsequent bacterial mineralization of an intermediate product(s) (Fig. 4, Table 5). Ring cleavage is postulated to occur as a result of microbial enzymes that denitrate CL-20, thus causing the cage to become destabilized (Bhushan et al., 2003, 2004a, 2004b). Mineralization of formate is a possible source of the 14CO2 that we observed, although glyoxal and other unidentified metabolites may also have been mineralized. Formate has been detected as an intermediate product of CL-20 degradation by every abiotic and biotic mechanism studied to date (Balakrishnan et al., 2003; Bhushan et al., 2004a, 2004b; Hawari et al., 2004; Monteil-Rivera et al., 2004; Szecsody et al., 2004). Glyoxal has only recently been detected as an intermediate product by reactions with nitroreductase (Bhushan et al., 2004a) or Fe0 (Balakrishnan et al., 2004). Based on the maximum amounts of 14CO2 produced, 2.4 to 3.3 C atoms per CL-20 molecule were converted to 14CO2. This molar ratio is slightly higher than the molar ratios of formate (0.49–2.0) measured in the above studies, indicating that formate did not supply all of the 14C converted to 14CO2. Since the expected ratio is 2 moles of formate and 2 moles of glyoxal per mole of CL-20 (Bhushan et al., 2004a), mineralization of glyoxal might explain the additional 14CO2 that we observed. Glyoxal is a natural product of cellular metabolism and it can be degraded by bacteria and fungi (Kersten 1990; Sakai et al., 2001; Whittaker et al., 1999). The lag phase and the slow rate of production of 14CO2 even after the parent CL-20 molecule had disappeared implied that a period of microbial adaptation to an intermediate product(s) was needed and that mineralization of the intermediate product(s) proceeded slowly. This may indicate that stimulation of other bacterial populations and/or enzymatic systems by the intermediate product(s) was necessary. The remaining 14C atoms in the mass balance were distributed among unidentified products that were either associated with the soil or dissolved in the aqueous phase.

In summary, abiotic and biological degradation mechanisms were important factors governing the fate of CL-20 in aerobic surface and subsurface soils. Based on our results it appears that biological degradation will have a greater influence on the fate of CL-20 in soils than abiotic degradation mechanisms when a readily available source of C is available. The biological and abiotic degradation rates of CL-20 suggest that CL-20 will not persist in these soils and thus will have a low potential to migrate to and within ground water at these sites. Mineralization of 14C–CL-20 to 14CO2 proved that ring cleavage occurred, since all carbons are part of the heterocyclic cage. In contrast, only biological degradation mechanisms were important to the fate of RDX in these same soils incubated under aerobic conditions. The presence of a microbial population that could degrade RDX under aerobic conditions was a key factor in determining the recalcitrance of this explosive in these soils. Differences in the rates of CL-20 and RDX biodegradation in the same soil suggest that the fate of each explosive in a soil must be evaluated on an individual basis.


   ACKNOWLEDGMENTS
 
This project was supported in part by grants from the Strategic Environment Research and Development Program (SERDP) and the U.S. Army Corps of Engineers Environmental Quality Technology research program. We thank Tonya Acuff and Margaret Richmond for technical assistance.


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 ABSTRACT
 INTRODUCTION
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 RESULTS
 DISCUSSION
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