EXTRACELLULAR ELECTRON SHUTTLE MEDIATED BIODEGRADATION OF HEXAHYDRO 1,3,5 TRINITRO 1,3,5 TRIAZINE BY MAN JAE KWON B.S., Korea University, 1999 M.S., Korea University, 2001 DISSERTATION Submitted in partial fulfillment of the requirements for the degree of Doctor of Philosophy in Environmental Engineering in Civil Engineering in the Graduate College of the University of Illinois at Urbana Champaign, 2009 Urbana, Illinois Doctoral Committee: Assistant Professor Kevin T. Finneran, Chair Professor Charles J. Werth Professor Joseph J. Stucki Assistant Professor Timothy J. Strathmann ABSTRACT EXTRACELLULAR ELECTRON SHUTTLE MEDIATED BIODEGRADATION OF HEXAHYDRO 1,3,5 TRINITRO 1,3,5 TRIAZINE Man Jae Kwon, Ph.D. Department of Civil and Environmental Engineering University of Illinois at Urbana Champaign, 2009 Kevin T. Finneran, Advisor Bioremediation is one suggested technology for hexahydro 1,3,5 trinitro
1,3,5 triazine (RDX), as well as alternate explosive contaminants. The potential use of extracellular electron shuttling compounds to mediate electron transfer between iron
(hydr)oxides and Fe(III) reducing microorganisms was investigated for decontaminating RDX. The ultimate goal of this study is to develop an optimal strategy for the biodegradation of nitramine explosive compounds.
This study demonstrated that Geobacter metallireducens , a model Fe(III) and electron shuttle reducing microorganism, utilize electron shuttles to stimulate cyclic nitramine reduction at rates significantly faster than those previously reported.
This work was the first to demonstrate electron shuttle mediated nitramine biodegradation with environmentally relevant pure cultures. Electron shuttle mediated explosives biodegradation was also investigated among four additional species of
Fe(III) and quinone reducing bacteria. All species showed the fastest RDX degradation when electron shuttling compounds were present although biotransformation kinetics were different among four Bacterial species. In addition, the critical issue for RDX remediation is how fast the ring cleavage metabolites are produced, and whether they are amenable to further degradation or ultimate mineralization. The higher production of desirable intermediate and mineralization compounds occurred with electron shuttling compounds and suggested that in situ
ii strategy can lead to more rapid and complete RDX mineralization by targeting
Fe(III) and electron shuttle reducing microorganisms and specific biotic abiotic reactions. RDX reduction directly stimulated by the electron shuttling compounds in
Fe(III) poor environments and indirectly stimulated by reactive Fe(II) in Fe(III) rich environments. This suggest that electron shuttle mediated RDX biodegradation is a reasonable strategy in both Fe(III) rich and Fe(III) poor environments.
Electron shuttling compounds also stimulated RDX mineralization in RDX contaminated aquifer material. The microbial communities associated with RDX biodegradation were characterized using amplified 16S rDNA restriction analysis.
The data demonstrated that a novel Fe(III) reducing community develops as RDX is degraded. This study also led to isolation and characterization of a novel species from
RDX contaminated aquifer material that is capable of reducing RDX. For in situ applications, different electron shuttling compounds were tested to identify an option that is both mechanistically feasible and cost efficient. G. metallireducens stimulated
RDX reduction with raw humics extract from commercial mulch and military smoke dye, suggesting that these quinone containing compounds could be an alternative source of electron shuttling compounds for in situ application.
iii ACKNOWLEDGMENTS This work would not have been possible without the support and encouragement of many people. I would like to express my deep and sincere gratitude to my advisor, Dr. Kevin T. Finneran, who quickly became my role model of a successful researcher in the field. He always waited for me with patience to explore unknown territories on my own. I have been very fortunate to have him as an advisor.
I would like to thank Dr. Timothy J. Strathmann who has been always there to listen and guide me with his knowledge of environmental organic chemistry. His understanding, encouragement and personal guidance have provided a good basis for the present thesis. Dr. Charles J. Werth has advised me with his profound insight in this field, steering my research in the right direction. His everyday life has taught me how to live as a scholar. I am so blessed to have had opportunities to meet him through my research and in classes. I am deeply grateful to Dr. Joseph W. Stucki for sharing his insight and experiences. He provided me an opportunity to understand microbial interaction with clay minerals.
I want to express my gratitude to Dr. Robert A. Sanford for sharing his time with me in discussing my research. His vast knowledge and his logical way of thinking have been of great value to me. I thank Scott R. Drew of GeoSyntec
Consultants for technical assistance and logistical management of field sampling and acquisition and for thoughtful discussions. I acknowledge Pam Sheehan of the
Picatinny Arsenal and Paul Hatzinger of Shaw Group for aquifer material sampling and preservation. I am also deeply indebted to the Finneran group members. I would like to thank Caitlin Bell, Yang Zhang, Jennifer Hatch, Kimberly Reinhaur, Kay
Dunett, Hossain Azam, Xiaofeng Ye, Patricia Shin, Wei Na, Annie Haluska, Brendan
Powers, and Jovan Popovic for their support and care, which have helped me
iv overcome setbacks and stay focused on my graduate studies. I also would like to thank Dr. Byung J. Kim, Manish Kumar, David Ladner, Marty Page, Tias Paul,
Lanhua Hu, Heather Martin, Dongwook Kim, Wonyoung Ahn, Sanghyung Lee,
Jinwoo Lee, Hongkyu Yoon, Changyong Zhang, and Youngchul Choi, for all their help, support, interest and valuable hints. Dr. Shaoying Qi manages all the labs and research orders on the 4 th floor. I feel that my research would have been much harder without his help. This work was supported by the Department of Defense Strategic
Environmental Research and Development Program (SERDP) project number ER
1377.
I cannot end without thanking my family, on whose constant encouragement and love I have relied throughout my time at the Academy. I would like to give my special thanks to Jisook whose patient love enabled me to complete this work.
v TABLE OF CONTENTS
1. INTRODUCTION TO BIODEGRADATION OF EXPLOSIVES ...... 1 AEROBIC BIODEGRADATION OF EXPLOSIVES...... 2 ANAEROBIC BIODEGRADATION OF EXPLOSIVES ...... 3 ELECTRON SHUTTLE MEDIATED BIODEGRADATION...... 4 CONCLUSIONS...... 5 2. MICROBIALLY MEDIATED BIODEGRADATION OF HEXAHYDRO 1,3,5 TRINITRO 1,3,5 TRIAZINE (RDX) BY EXTRACELLULAR ELECTRON SHUTTLING COMPOUNDS ...... 9 INTRODUCTION...... 9 MATERIALS AND METHODS...... 10 RESULTS ...... 12 DISCUSSION ...... 17 3. HEXAHYDRO 1,3,5 TRINITRO 1,3,5 TRIAZINE (RDX) AND OCTAHYDRO 1,3,5,7 TETRANITRO 1,3,5,7 TETRAZOCINE (HMX) BIODEGRADATION KINETICS AMONGST SEVERAL FE(III) REDUCING GENERA ...... 32 INTRODUCTION...... 32 MATERIALS AND METHODS...... 34 RESULTS ...... 36 DISCUSSION ...... 38 4. BIOTRANSFORMATION PRODUCTS AND MINERALIZATION POTENTIAL FOR HEXAHYDRO 1,3,5 TRINITRO 1,3,5 TRIAZINE (RDX) IN ABIOTIC VERSUS BIOLOGICAL DEGRADATION PATHWAYS WITH ANTHRAQUINONE 2,6 DISULFONATE (AQDS) AND GEOBACTER METALLIREDUCENS ...... 48 INTRODUCTION...... 48 MATERIALS AND METHODS...... 50 RESULTS ...... 53 DISCUSSION ...... 54 5. HEXAHYDRO 1,3,5 TRINITRO 1,3,5 TRIAZINE (RDX) REDUCTION IS CONCURRENTLY MEDIATED BY DIRECT ELECTRON TRANSFER FROM HYDROQUINONES AND THE RESULTING BIOGENIC FE(II) FORMED DURING ELECTRON SHUTTLE AMENDED BIODEGRADATION ...... 64
vi INTRODUCTION...... 64 MATERIALS AND METHODS...... 66 RESULTS AND DISCUSSION...... 71 CONCLUSIONS...... 79 6. HEXAHYDRO 1,3,5 TRINITRO 1,3,5 TRIAZINE (RDX) MINERALIZATION AND BIOLOGICAL PRODUCTION OF 4 NITRO 2,4 DIAZABUTANAL (NDAB) IN THE PRESENCE OF EXTRACELLULAR ELECTRON SHUTTLING COMPOUNDS AND A UNIQUE FE(III) REDUCING MICROBIAL COMMUNITY...... 88 INTRODUCTION...... 88 MATERIALS AND METHODS...... 91 RESULTS AND DISCUSSION...... 96 7. CHARACTERIZATION OF A HEXAHYDRO 1,3,5 TRINITRO 1,3,5 TRIAZINE (RDX) DEGRADING, NOVEL ANAEROBE ISOLATED FROM CONTAMINATED AQUIFER MATERIAL...... 118 SUMMARY ...... 118 INTRODUCTION...... 119 MATERIALS AND METHODS...... 119 RESULTS AND DISCUSSION...... 122 8. RAW HUMICS EXTRACT AND MILITARY SMOKE DYE AS EXTRACELLULAR ELECTRON SHUTTELS FOR BIODEGRADATION OF HEXAHYDRO 1,3,5 TRINITRO 1,3,5 TRIAZINE (RDX)...... 132 INTRODUCTION...... 132 MATERIALS AND METHODS...... 134 RESULTS ...... 137 DISCUSSION ...... 139 9. SUMMARY AND CONCLUSIONS...... 151 10. REFERENCES...... 159 11. APPENDIX A: CHEMICALS AND ANALYTICAL METHODS...... 176 12. CURRICULUM VITAE ...... 179
vii CHAPTER 1 INTRODUCTION TO BIODEGRADATION OF EXPLOSIVES
The cyclic nitramine explosives hexahydro 1,3,5 trinitro 1,3,5 triazine (RDX) and octahydro 1,3,5,7 tetranitro 1,3,5,7 tetrazocine (HMX) are contaminants at many military sites and live fire training installations (Spalding and Fulton, 1988; Haas et al., 1990; Hawari and Halasz, 2002). RDX (60 mg/L [270 M]) and HMX (5 mg/L
[17 M]) solubility in water is relatively low (Talmage et al., 1999); however, once dissolved they can migrate rapidly through the vadose and saturated zones to contaminate down gradient aquifers. Groundwater contaminated by these explosives is of growing environmental concern because of human health effects of RDX and
HMX. RDX is a possible human carcinogen (lifetime health advisory for exposure to
RDX in drinking water: 2 g/L [9 nM]), and HMX damages the central nervous system (lifetime health advisory for exposure to HMX in drinking water: 0.4 mg/L
[1.4 M]) (Lynch, 1988; Lachance et al., 1999). Despite several reports suggesting that these compounds are readily biodegraded, RDX and HMX persist in subsurface environments (Meyers et al., 2007). A recent study reported that RDX remains one of the most significant contaminants at the Camp Edwards Military Reservation despite several years of remediation efforts (Clausen et al., 2004). RDX contamination is a significant issue for the department of defense (DoD). Recent press at the
Massachusetts Military Reservation (MMR) located in Sandwich, MA, reported that
RDX was identified in a plume moving offsite at a mean concentration of 290 g/L
(highest 370 g/L), which is over 100 times the maximum contaminant level (MCL)
(Wani et al., 2002).
1 Few cost effective technologies exist for the treatment of RDX compounds in groundwater. Pump and treat systems are typically inefficient for the remediation of groundwater plumes largely because they do not address the source of contamination, and because of the large volumes of groundwater that must be treated for hydraulic control and regulatory compliance (Abdelouas et al., 1998). In addition, these strategies merely remove the groundwater and transfer the contaminants to another medium rather than degrading the contaminants. Permeable reactive barriers (PRB) employing zero valent iron (ZVI) show promise in reducing and/or treating nitramine compounds, but PRB installation is known to be technically and economically infeasible at sites with deep or wide plumes (Wolbarst et al., 1999).
Bioremediation is one alternative for explosives contamination and several mechanisms have been proposed using contaminated sediment or soil or pure and mixed cultures (Adrian et al., 2003; Adrian and Arnett, 2004; Sherburne et al., 2005;
Thompson et al., 2005; Meyers et al., 2007). However, the reported degradation rates of cyclic nitramines have varied (Hawari et al., 2000b; Zhao et al., 2002; Wani and
Davis, 2003; Wani and Davis, 2006).
AEROBIC BIODEGRADATION OF EXPLOSIVES
RDX or HMX has a cyclic, nitrogen containing molecular structure and they are moderately resistant to aerobic biodegradation (Figure 1.1)(Bradley and Chapelle,
1995; Weissmahr et al., 1999). RDX will be biodegraded aerobically in the presence of specific microorganisms (Table 1.1), but specialized enrichment conditions are required for these organisms to proliferate (Bradley and Chapelle, 1995), and several reports suggest limited degradation kinetics (Yinon, 1990; Harvey et al., 1991).
RDX and HMX are utilized as nitrogen or carbon source for aerobic microorganisms.
Nitrite formation following initial enzymatic attack was frequently reported (Table 2 1.1). Aerobic bioremediation is possible in shallow soil or groundwater that has sufficient oxygen, but is technically difficult in groundwater that has become anaerobic. Environments with RDX contamination may be co contaminated with compounds that promote anaerobic conditions.
ANAEROBIC BIODEGRADATION OF EXPLOSIVES
To date, most bioremediation of explosives contamination were investigated under anaerobic conditions (Table 1.1) because subsurface RDX or HMX plumes are mostly anaerobic and these compounds degrade more readily under anaerobic conditions (Garg et al., 1991 ; Adrian and Arnett, 2004; Zhang and Bennett, 2005).
Similar to aerobic strategy, several anaerobic microorganisms utilized explosives as a nitrogen source. Anaerobic biodegradation through the nitro nitroso degradation pathway (initial reductive step) has also been suggested as an effective strategy for
RDX and HMX degradation (Adrian et al., 2003). Transformation of RDX through mono , di , and tri nitroso products has been proposed (McCormick et al., 1981) and such reduction leads to destabilization, ring cleavage, and mineralization of RDX under anaerobic conditions. Degradation intermediates are susceptible to aerobic as well as anaerobic mineralization; however, mineralization is nearly an order of magnitude greater under anaerobic conditions. Direct microbial RDX or HMX reduction has been reported (Zhao et al., 2002; Zhao et al., 2003b), and this process can be catalyzed by microorganisms that specifically transfer electrons to nitramine compounds (whether the reduction is co metabolic or linked to energy generation is still debated); however, without these microorganisms, the overall biodegradation kinetics will be limited. Alternate degradation pathways are possible, but are less easily stimulated in subsurface environments (Pennington and Brannon, 2002;
Crocker et al., 2006). 3 ELECTRON SHUTTLE MEDIATED BIODEGRADATION
Electron transfer from hydroquinones (Figure 1.1)(the functional electron transfer unit of HS) (Scott et al., 1998) to nitro functional groups has been reported.
Juglone and lawsone transfer electrons to nitroaromatic compounds with a bulk chemical reductant providing the electrons for the quinones (Schwarzenbach et al.,
1990). Electron shuttling from microorganisms is more relevant to in situ remediation strategies since native microbial communities provide the necessary reducing power, eliminating the need for a bulk reductant within the system (Schwarzenbach et al.,
1990). In addition, extracellular electron shuttling compounds have been proposed as an electron transfer mediator between microorganisms and solid phase minerals to stimulate iron reduction by eliminating the need for cell oxide contact (Figure
3)(Lovley et al., 1996b). Electron shuttle mediated contaminant transformation has also been demonstrated for the BTEX compounds (Lovley et al., 1994; Cervantes et al., 2001; Finneran and Lovley, 2001), methyl tert butyl ether (MTBE) 1 (Finneran and
Lovley, 2001; Finneran et al., 2001), carbon tetrachloride (Cervantes et al., 2004;
Doong and Chiang, 2005), chlorinated compounds such as dicholoroethene and vinyl chloride (Bradley et al., 1998), and metals such as uranium (Fredrickson et al., 2000a;
Finneran et al., 2001; Lloyd and Lovley, 2001; Holmes et al., 2002; Jeon et al., 2004).
More recently humic substances (HS), and the HS analog anthraquinone 2,6 disulfonate (AQDS) have been investigated for their role in reductive transformation of explosive compounds (Borch et al., 2005). Trinitrotoluene (TNT) reduction was accelerated by adding AQDS to fermentative cultures (Borch et al., 2005).
Fermentative electron transfer to Fe(III) has been previously reported, and these data suggest that this non respiratory electron transfer mechanisms may also stimulate 1 BTEX and MTBE are the electron donor to the cell and these compounds do not interact with electron shuttles
4 AQDS mediated electron shuttling to TNT. Fermentative AQDS/Fe(III) reduction is a fortuitous reaction from which the cells derive no energetic benefit; therefore the rate and extent of these reactions will differ greatly from population to population
(Lovley, 1991). These data are promising; however, it is unclear whether the specific fermentative organisms utilized within these experiments are competitive in environments typically contaminated with TNT or other explosive compounds (Borch et al., 2005). Given that fermentative electron transfer to Fe(III) and/or AQDS is non specific, it is more feasible to promote the reactions for organisms that specifically reduce these compounds (i.e., quinones) to gain energy for growth. An alternate study with marine sediment enrichments suggested that Shewanella species were the predominant phylotypes within the most active RDX reducing enrichment.
Shewanella are known Fe(III) reducing bacteria; however it was not clear whether reactive Fe(II) within the enrichments was responsible for any part of the reaction pathway (Zhao et al., 2004b; Zhao et al., 2004c).
CONCLUSIONS Anaerobic biodegradation of nitramine compounds is more likely to be applied to in situ because of mechanical and economical reason. Electron shuttling compounds are known to stimulate contaminant reduction. Since electron shuttling compounds are catalytic, in situ applications may only require a low concentration of these compounds to stimulate contaminant reduction. In addition, Fe(III) and electron shuttle (e.g., humic substance) reducing microorganisms have been identified in shallow and deep aquifer material, freshwater (Lovley and Phillips, 1986) and marine sediment (Hayes et al., 1999), soil (Coates et al., 1995), and extreme environments such as hot springs and volcanic sediment (Kashefi and Lovley, 2003). The ubiquity of Fe(III) and electron shuttling compound reducing microorganisms increases the
5 likelihood that remediation strategies predicated on their physiology will be successful in many subsurface environments (Coates et al., 1998).
6
2H + + 2e -
Figure 1.1. Structure of RDX, HMX, oxidized quinone, and reduced quinone
7 t ; o a i n r e = t c a / a n b a g QRB = quinone reducin d ble biochemical mechanisms for degradation; DIRB = dissimilatory iron reducing bacteria; c NT = not tested; b SRB = sulfate reducing bacteria Table 1.1. RDX or HMX degrading bacteria and possi available; e
8 CHAPTER 2
MICROBIALLY MEDIATED BIODEGRADATION OF HEXAHYDRO 1,3,5
TRINITRO 1,3,5 TRIAZINE (RDX) BY EXTRACELLULAR ELECTRON
SHUTTLING COMPOUNDS 2
INTRODUCTION
Most anaerobic strategies for RDX to date have focused on direct microbial reduction of the nitro functional groups on the cyclic structure as the sole terminal electron acceptor (Zhao et al., 2004c). This strategy is effective only when specific microorganisms that respire nitramine compounds are widespread in subsurface systems. In the absence of these microorganisms the reactions may be slow, limiting this strategy in many environments.
Extracellular electron shuttling may be one approach for cyclic nitramines.
However, electron shuttle mediated degradation of cyclic nitramines, including RDX, has not been reported yet. Extracellular electron shuttling encompasses all reactions that are catalyzed by microbial reduction of the shuttles – whether it is direct interaction with the reduced shuttle or with Fe(II) resulting from the reaction. Fe(III) and humic substance (HS) reducing microorganisms have been identified in shallow and deep aquifer material, freshwater (Lovley and Phillips, 1986) and marine sediment (Hayes et al., 1999), soil (Coates et al., 1995), and extreme environments such as hot springs and volcanic sediment (Kashefi and Lovley, 2003). A few recent reports (Gregory et al., 2004; Williams et al., 2005) suggest that RDX was transformed by reactive Fe(II), the product of Fe(III) respiration. The ubiquity of
Fe(III) and HS reducing microorganisms increases the likelihood that remediation 2 Reproduced in part with permission from Kwon, M.J. and K.T. Finneran. Applied and Environmental Microbiology . 2006, 72 : 5933 – 5941
9 strategies predicated on their physiology will be successful in many subsurface environments (Coates et al., 1998).
In this study, indirect or direct electron transfer to RDX via reduced purified
HS, reduced anthrquinone 2,6 disulfonate (AQDS), and Fe(II) was investigated.
Understanding electron transfer mechanisms via an electron shuttle to RDX (versus direct electron transfer from microbial respiration) will be important to establish strategies for groundwater bioremediation. The objectives of this study were to 1.) verify specific mechanisms for RDX reduction by Fe(III) reducing microorganisms and HS, and 2.) to determine if stimulating Fe(III) and HS reduction will increase the rate and extent of RDX biodegradation. In addition, the study was designed to determine if HS mediated RDX transformation is microorganism specific or a general phenomenon among different HS reducing species.
MATERIALS AND METHODS Medium and culturing conditions . The basal medium consisted of (g l 1 unless specified otherwise): NaHCO 3 (2.5), NH 4Cl (0.25), NaH 2PO 4 H 2O (0.6), KCl (0.1), modified Wolfe’s vitamin and mineral mixtures (each 10 ml l 1) and 1 mL of 1mM
Na 2SeO 4. Electron acceptors used with the medium included soluble Fe(III) citrate
(45 mM), poorly crystalline Fe(III) (hydr)oxide (50 mmol/L), or AQDS (5 mM). All cultures were maintained as previously described (Finneran et al., 2003).
Pure phase incubations . Chemically reduced AQDS (C AH 2QDS) was prepared by sparging the medium bottle with H 2:CO 2 (80:20, vol/vol) in the presence of palladium covered alumina pellets as previously described (Lovley et al., 1999). The chemically reduced AQDS was filtered through a 0.2 m sterilized PTFE filter into a pre sterilized, anaerobic serum bottle. Biologically reduced AQDS (B AH 2QDS) was prepared by incubating Geobacter metallireducens in AQDS medium (5 mM) . The
10 B AH 2QDS was filtered through a 0.2 m sterilized PTFE filter into a pre sterilized, anaerobic serum bottle to remove cells. Biologically reduced Fe(II) was prepared in a similar manner using 45 mM Fe(III) citrate medium in lieu of AQDS medium.
The concentrations of reduced AQDS (chemical or biological AH 2QDS) tested were 500, 300, and 100 M for the incubations with RDX; 300, 100, and 50 M for the incubations with MNX; 200, 100, and 50 M for the incubations with DNX.
Incubations were performed at 25°C. The starting RDX concentration was either 100
M or 50 M, depending on the specific experimental conditions.
Experiments were initiated by injecting RDX into the buffer with AH 2QDS.
Samples were collected periodically via anaerobic syringe and needle; samples were filtered through sterile, 0.2 m PTFE filters prior to analyses. RDX was quantified at each time point. AH 2QDS concentration was quantified for AH 2QDS/ AQDS incubations and Fe(II) was quantified for ferrous iron incubations.
Resting cell suspension incubations . Cell cultures ( G. metallireducens or G. sulfurreducens) were grown in freshwater medium with acetate as the sole electron donor and Fe(III) citrate as the sole terminal electron acceptor. One liter of cell culture was harvested during logarithmic growth phase and centrifuged (at 5000 RPM for 15 minutes) to form a dense cell pellet and cell suspensions were conducted as previously described (Nevin and Lovley, 2002b).
Electron acceptors were amended from concentrated, anaerobic, sterile stock solutions. Electron acceptors incubated with the cells including humic acids (0.25 g/L), poorly crystalline Fe(III) oxide (45 mM), AQDS (5 mM), humic acids plus poorly crystalline Fe(III) oxide, and AQDS plus poorly crystalline Fe(III) oxide.
Cells were incubated at 30°C. Acetate was added at a final concentration of 20 mM to each suspension as the sole electron donor. An aliquot (0.3 ml) of the resting cells was added to the sealed pressure tubes to initiate each experiment. 11 Samples were collected periodically via anaerobic syringe and needle. RDX was quantified at each time point. AH 2QDS concentration was quantified for AQDS incubations and Fe(II) was quantified for ferrous iron incubations. Cell protein normalized decay rates were calculated based on pseudo first order degradation coefficients.
RESULTS
RDX reduction by biologically and chemically reduced AH 2QDS . In order to evaluate if reduced HS transfer electrons to the cyclic nitramine RDX, the HS analog
AQDS (Figure 2.1) was biologically and chemically reduced, and incubated in cell free, anaerobic mixtures with RDX. AQDS was used in lieu of purified HS because it is a defined molecule (molecular mass: 366.32 g/mol) that has been well characterized with respect to its electron shuttling capacity. In addition, its color changes from opaque pink to vivid orange as it changes from oxidized to reduced state; therefore it can be quantified spectrophotometrically at 450 nm. Two different concentrations of chemically and biologically reduced AQDS were amended to 50 M of RDX to determine if the reduction rate of RDX was different with the reduced AQDS produced by different methods. The quantified reduction rate of RDX between chemically and biologically reduced AQDS was not significantly different; therefore, subsequent pure phase incubations were performed with only biologically reduced
AQDS.
RDX is reduced by a series of stepwise 2 electron transfers generating nitroso intermediates (Figure 2.1). Stoichiometry of electron transfer between reduced
AQDS and RDX was investigated with three different concentrations of microbially reduced AQDS (Figure 2.2). 100 M RDX was completely transformed within 9 hours in the incubations when amended with 500 M and 300 M of B AH 2QDS. 12 However, RDX was reduced to approximately 61 M with 100 M of B AH 2QDS
(Figure 2.2A). This result was within range of the expected stoichiometry for oxidation of AH 2QDS coupled to RDX reduction (with TNX as the end product of the reductive pathway). Approximately one third of the 100 M RDX was reduced when one third of the stoichiometric equivalent of AH 2QDS was provided, suggesting that
RDX was reduced to TNX, with TNX removed by an unidentified reaction pathway.
Reduced AQDS transfers two electrons per mole in the coupled oxidation/reduction reaction and RDX accepts six electrons per mole (to TNX.) The half reactions, and full reaction, are as follows:
+ C3H6N6O6 (RDX) + 6H + 6e C 3H6N6O3 (TNX) + 3H 2O (Red. half reaction)
+ AH 2QDS AQDS + 2H + 2e (Ox. half reaction)
C3H6N6O6 (RDX) + 3AH 2QDS C 3H6N6O3 (TNX) + 3AQDS + 3H 2O (Full reaction)
Therefore, three times as much B AH 2QDS (molar basis) is needed to completely reduce RDX. The two likely reaction pathways would be RDX MNX, in which case the AH 2QDS would have been oxidized with stoichiometric MNX accumulation, or RDX TNX, in which case AH 2QDS would be oxidized without MNX or DNX accumulation. MNX, DNX, and TNX did not accumulate; therefore the second pathway is the likely pathway. Lack of TNX accumulation will be discussed below.
The concentration of AH 2QDS decreased coupled to RDX reduction. 500 M
AH 2QDS was oxidized to approximately 200 M and 300 M AH 2QDS was oxidized to approximately 50 M concomitantly with RDX reduction (Figure 2.2B). When
AH 2QDS (100 M) was limiting relative to RDX (at 100 M), it was completely oxidized within 8 hours. 13 MNX and DNX reduction by biologically reduced AH 2QDS. The metabolites hexahydro 1 nitroso 3,5 dinitro 1,3,5 triazine (MNX), hexahydro 1,3 dinitroso 5 nitro 1,3,5 triazine (DNX), and hexahydro 1,3,5 trinitroso 1,3,5 triazine (TNX) did not significantly accumulate during AH 2QDS mediated RDX reduction; therefore the individual metabolites were incubated with AH 2QDS to determine if they can be directly reduced. 50 M MNX was directly reduced by AH 2QDS at all AH 2QDS concentrations tested; however, the extent of reduction differed among the treatments
(Figure 2.3A). Excess AH 2QDS (300 M) completely reduced MNX below detection limits within 8 hours. When the stoichiometric equivalent of AH 2QDS (100 M) was provided, reduction was slower and did not proceed to completion. However, previous experiments suggested that the stoichiometric equivalent completely reduced
MNX in a much shorter time frame (Figure 2.3C). 50 M AH 2QDS reduced approximately 20 M of MNX (Figure 2.3A), which is within range of the expected stoichiometry.
DNX was also reduced by AH 2QDS; however, the DNX stock was not chemically pure and the reaction did not closely match the expected stoichiometry for
DNX (Figure 2.3B). The original stock was composed of approximately 58% DNX,
34% MNX, and 8% TNX. AH 2QDS in excess of the necessary stoichiometry (200
M) completely reduced the mixture of DNX plus MNX in approximately 30 hours.
However, past experiments suggest that DNX is reduced in less than 4 hours (Figure
2.3D). MNX and DNX did not accumulate to a significant extent in any AH 2QDS amended experiments with RDX as the starting material. TNX was transformed as well (data not shown) by an unidentified pathway; the mechanism is currently under investigation using uniformly radiolabeled [ 14 C] RDX. RDX reduction kinetics slowed at 4 °C (Table 2.1). Although MNX and TNX were quantifiable neither of the
14 nitroso intermediates accumulated in the incubations. DNX was still not quantified at this temperature.
RDX reduction by biologically reduced, soluble Fe(II). 50 M RDX was reduced to 19 M and 39 M by 1.2 mM and 600 M of soluble Fe(II), respectively (Table 1).
RDX reduction by Fe(III)/extracellular electron shuttles in the presence of
Fe(III) reducing cells ( Geobacter metallireducens ). RDX was reduced in all experimental incubations containing resting cells; however, the rate and extent of
RDX reduction differed greatly among treatments. AQDS stimulated the fastest rate of RDX reduction (Figure 2.4A). RDX was reduced to below detection limits in less than 12 hours regardless of the presence of acetate. The initial degradation rate of
RDX with the presence of acetate was faster than that of RDX without acetate. Cells of G. metallireducens without AQDS also reduced RDX; RDX was still detectable at
12 hours in cells only incubations. These are the first data demonstrating RDX reduction by cells within the family Geobacteraceae . Later time points demonstrated that cells alone reduced RDX in approximately 50 hours, which was comparable to purified humic substance mediated RDX reduction. The maximum concentration of
MNX quantified during the incubations with AQDS was 0.25 M and it was no longer detected within 12 hours (Figure 2.4B). RDX was completely reduced to concentrations below detection within 50 hours with purified humic substances
(Figure 2.4C). MNX transiently accumulated to 0.8 M in humics substance amended incubations, and quickly decreased at approximately 20 hours (Figure 2.4D). RDX reduction was the slowest with poorly crystalline Fe(III) oxide as the sole electron acceptor. RDX slowly decreased from 40 M to 30 M with poorly crystalline Fe(III) oxide for 100 hours (Figure 2.4E). MNX accumulated in these incubations, although 15 the concentration was not a stoichiometric increase suggesting partial MNX reduction
(Figure 2.4F). MNX concentration eventually decreased below the detection limit with longer incubation times. DNX and TNX did not accumulate significantly within the time frame of these experiments. Heat killed cells did not reduce RDX in the presence or absence of acetate.
RDX reduction was slow with only poorly crystalline Fe(III) oxide (Figure
2.4E). Adding humic acids or AQDS to Fe(III) containing incubations increased the reduction rates by approximately 5 and 66 times, respectively (Figure 2.5A and 2.5B).
These rates of RDX degradation were calculated from the pseudo first order RDX decay constants (Table 2.1). Fe(II) concentrations were higher in AQDS and HS amended incubations than poorly crystalline Fe(III) oxide alone (Figure 2.5C). RDX was not reduced in the absence of cells; therefore only the humic amended “no cells” control is presented (results with AQDS were identical.)
RDX reduction by Fe(III)/extracellular electron shuttles in the presence of
Fe(III) reducing cells ( Geobacter sulfurreducens ). Similar results were obtained with resting cell suspensions of G. sulfurreducens , another member of the
Geobacteraceae which has variant physiological properties from G. metallireducens
(Lovley et al., 1993; Caccavo et al., 1994). RDX was reduced with cell suspensions of
G. sulfurreducens with several electron acceptors including AQDS, HS, and poorly crystalline Fe(III) oxide. Direct electron transfer from AQDS or HS was faster than
Fe(III) Fe(II) mediated electron transfer to RDX. All G. sulfurreducens results are summarized as individual reaction rates in Table 2.1.
16 DISCUSSION These results suggest that electron shuttle mediated RDX reduction is favorable and that RDX is reduced faster by extracellular electron shuttles, including
HS and the HS analog AQDS, than direct microbial reduction or bound Fe(II). HS reduction has been identified in numerous subsurface environments and is catalyzed by microorganisms that are ubiquitous in their distribution (Coates et al., 1998;
Lovley et al., 1998; Scott et al., 1998). Direct reduction of RDX or its nitroso metabolites has been reported, but results suggest that activity may differ among contaminated environments (Bhushan et al., 2002; Wani and Davis, 2003; Zhao et al.,
2004c). Effective remediation strategies for explosive residue contaminants such as the cyclic nitramines (RDX and HMX) must be effective in many environments before they will be accepted by the regulatory community.
RDX biodegradation has been reported for aerobic and anaerobic systems; anaerobic biodegradation is the preferential pathway because the nitro groups must be reduced prior to ring cleavage and mineralization to CO 2 under anaerobic or aerobic conditions (Hawari et al., 2000a; Wani and Davis, 2003). Recent data demonstrated that aerobic, methylotrophic bacteria will degrade a ring cleavage product of RDX –
4 nitro 2,4 diazabutanal (Fournier et al., 2005); however, the ring must first be cleaved for this reaction, which is faster under anaerobic conditions. All reported aerobic data are promising, but RDX contaminates anaerobic subsurface and sedimentary environments in which aerobic metabolism is limited and inefficient
(Weissmahr et al., 1999; Wolbarst et al., 1999). In addition, given the extent of some
RDX plumes (Wani et al., 2002), no strategy should be overlooked because one plume may encompass several geochemical zones.
Strategies which promote RDX reduction through MNXDNX TNX have been investigated in more detail than alternative mechanisms (Hawari et al., 2000a;
17 Oh et al., 2001). RDX biodegradation has been reported for environmental samples
(e.g. marine sediment), mixed anaerobic cultures, or pure cultures (Oh et al., 2001;
Zhao et al., 2002; Zhao et al., 2005). Mixed enrichment cultures cultivated from marine sediment reduced RDX through its nitroso intermediates and ultimately led to mineralization of nearly 60% of the RDX in one series of incubations (Zhao et al.,
2004b; Zhao et al., 2005). The mineralization activity was not consistent, and alternate batches yielded different results (Zhao et al., 2004b). However, Shewanella species were the dominant microbial population in the most active enrichment, which suggests that in situ Fe(III) reducing microorganisms may be responsible for RDX biodegradation. One novel species within the genus Shewanella was isolated from this sediment, which suggests that Fe(III) reduction may have been the dominant terminal electron accepting process (TEAP) (Zhao et al., 2004b). This previous investigation did not explore the link between Fe(III) reduction and RDX transformation, however, and the authors suggested that direct RDX reduction is the primary mechanism. While the data presented above demonstrate that
Geobacteraceae will directly reduce RDX, it is transformed faster by abiotic electron shuttling from reduced quinones. Reports of RDX biodegradation have been sporadic, and no published data have yet to suggest a link between a specific group of microorganisms and RDX reduction. The inconsistent degradation patterns may be due to environmental samples with different dominant TEAPs promoting a wide array of degradation kinetics. However, the most consistent RDX reduction data prior to those reported here have been in the presence of Fe(III) reducing bacteria (Williams et al., 2005).
Gregory et al demonstrated that reactive biogenic Fe(II) (soluble ferrous iron adsorbed to the surface of magnetite) reduced RDX through its nitroso intermediates
(Gregory et al., 2004). The metabolites accumulated for several days prior to their 18 removal, and TNX accumulated in most bottles (Gregory et al., 2004). The data presented above with Fe(III) agree with these past studies suggesting that Fe(II) will transform RDX; however, soluble Fe(II) reduced RDX much more slowly than Fe(II) generated from poorly crystalline Fe(III) oxide (i.e. magnetite). The past studies did not, however, stimulate Fe(III) reduction by adding soluble electron shuttles. As described above, processes that stimulate the rate of Fe(III) reduction also stimulate the rate of RDX reduction. In all cases HS reduced RDX and its intermediates faster than biogenic Fe(II) (with reaction rates on the order of hours rather than days)
(Figures 2.3 – 2.5: Table 2.1).
HS have been reported to facilitate degradation of several organic contaminants (Stevenson, 1982; Lovley et al., 1996b) as an electron acceptor or shuttle to Fe(II) which promotes oxidation of an organic molecule such as benzene, toluene, or methyl tert butyl ether (MTBE) (Lovley et al., 1996a; Finneran and
Lovley, 2001; Finneran et al., 2001). HS and quinone analogs such as AQDS, juglone, and lawsone also reduce inorganic and organic molecules (Schwarzenbach et al., 1990; Lovley et al., 1998; Schwarzenbach et al., 2003). AQDS was reported to reduce U(VI) in solution or adsorbed to the surface of oxide solids (Fredrickson et al.,
2000b; Jeon et al., 2004). Juglone and lawsone reduced trinitrotoluene (TNT) in aqueous suspension (Schwarzenbach et al., 1990). Although high molecular weight
HS are insoluble and can merely adsorb organic contaminants, low molecular weight
HS are soluble and will likely promote these reactions in situ. Remediation strategies predicated on low molecular weight HS are relatively benign in that the HS are catalytic and only a low concentration is needed to promote these reactions (Lovley et al., 1998).
Microorganisms within the family Geobacteraceae are generally reported as model Fe(III) reducing Bacteria (Lovley et al., 1991; Coates et al., 1996). They have 19 been identified in pristine and contaminated subsurface and sedimentary environments, and are implicated in iron and HS mediated processes – including contaminant transformation – in both pure culture and in situ (Coates et al., 1996;
Coates et al., 1998). A large body of data exists for their physiology, which is why two species within the Geobacter genus were selected to first test the HS RDX biodegradation pathway. The data presented herein are the first demonstrating cyclic nitramine reduction by a member of the Geobacteraceae ; however, the electron shuttle mediated mechanism was significantly faster than direct electron transfer from high density resting cells. In addition, the two species tested did not conserve energy for growth coupled to RDX reduction (data not shown.) Numerous genera are reported to reduce Fe(III), AQDS, or natural HS; alternative genera that reduce HS include Desulfitobacterium , Anaeromyxobacter , Shewanella , and Rhodoferax (Lovley et al., 1998; Finneran et al., 2003; He and Sanford, 2003). Initial experiments with the genera mentioned here suggest that all of these organisms promote the same reactions (data not shown.) It is the ubiquity of Fe(III) and HS reduction (and the diversity of the microorganisms involved) that makes the proposed bioremediation strategy an attractive alternative to RDX reduction by direct cellular electron transfer processes.
The data presented here are the first demonstrating RDX reduction by reduced
HS or the reduced HS analog AH 2QDS. Reduction rates were similar for chemically reduced or biologically reduced AQDS; however the eventual strategy is based on microbial HS reduction therefore the biological mechanism is more relevant to the ongoing investigation. Past studies suggest RDX reduction on the orders of days, weeks, or months. AH 2QDS and HS mediated RDX reduction was significantly faster, on the order of hours for complete degradation. Although the number of reactive pathways for reduced quinones will increase in environmental media, these 20 data suggest that HS and Fe(III) reduction stimulated by microorganisms will attenuate RDX in situ.
RDX is reduced through three nitroso metabolites; remediation strategies must attenuate these compounds as well as the parent compound because the mono and di nitroso form (MNX and DNX, respectively) are also relatively stable and pose a groundwater hazard. TNX, the trinitroso metabolite, is less stable and may degrade via several putative pathways in situ (Larson et al., 2005). However, several reports have demonstrated that TNX may accumulate during RDX biodegradation
(McCormick et al., 1981; Oh et al., 2001; Sheremata et al., 2001), which suggests that the ultimate fate of the nitroso metabolites depends on the site conditions. Initial experiments with MNX and DNX suggested reduction in less than two hours (Figures
2.3C and 2.3D). However, experimental conditions were different (higher temperature, lower pH) which may have altered reduction kinetics. Later experiments
(Figures 2.3A and 2.3B) run at more biologically relevant conditions demonstrated
MNX and DNX reduction over 8 12 hours, which is still a reasonable time frame for bioremediation.
TNX was not recovered at stoichiometric concentration in any of the incubations within this research, suggesting that TNX in either: a) unstable and degrades auto catalytically, or b) degraded by an as yet unidentified reductive mechanism. These HS and AQDS mediated reactions were pH and temperature specific. As expected, lower temperature decreased the rate of reaction with RDX and all intermediates (Table 2.1). Small fluctuations in the pH ( ± 0.2 pH units) did not alter the reaction stoichiometry or rate. Larger pH fluctuations may change the reaction rate and will be tested.
RDX reduction kinetics did not change significantly in the presence of resting cell mass with either Geobacter metallireducens or G. sulfurreducens . Cell free 21 AH 2QDS completely reduced RDX in approximately 8 hours; both AQDS amended cell suspensions reduced RDX completely within 12 hours. However, this does suggest that the rate limiting step is likely microbial AQDS reduction rather than abiotic electron transfer from AH 2QDS to RDX. Cells reduced RDX in the presence of AQDS whether acetate was added as an electron donor or not. This has been previously described for several metal reducing microorganisms including Thermus strain SA and Deinococcus radiodurans R1 (Kieft et al., 1999; Fredrickson et al.,
2000a). It has been referred to as “endogenous respiration” and refers to the cellular capacity to utilize electron donors from cell mass (either storage molecules or cellular lysate debris) to continue metabolism. Given that Geobacteraceae are not reported to generate storage molecules under these conditions it is likely that the acetate free respiration was promoted by cellular debris. It was not non specific electron transfer from reduced cytochromes (or equivalent electron transfer molecules) to AQDS or HS because air oxidized cell suspensions behaved in a similar manner with RDX being reduced in the absence of acetate (Table 2.1).
Natural HS also stimulated RDX reduction in the presence of both cell suspensions. HS mediated RDX reduction was slower than AQDS mediated RDX reduction (Figure 2.3 2.5; Table 2.1). This is not surprising given that AQDS is a small molecule with a single, easily accessible (i.e. not sterically hindered) quinone moiety that transfers two electrons per mole (Lovley et al., 1996b). Natural HS are large, undefined molecules with varying quinone content, and each individual quinone functional group may be slightly less accessible (for either cell quinone interaction or hydroquinone RDX interaction) than the comparable AQDS or AH 2QDS reaction mechanisms. Previous investigations with electron shuttling demonstrated that
AQDS, natural (Aldrich) humics, and purified International Humic Substance Society
(IHSS) humics shuttle electrons at different rates to Fe(III) and other electron 22 acceptors (Lovley et al., 1998; Nevin and Lovley, 2002a). Follow up experiments with IHSS humics will be performed. However, these data suggest that different HS
(or analogs/quinones) will reduce RDX at different rates with slightly different metabolite accumulation dynamics, and that for in situ applications different HS sources must be tested to identify an option that is both mechanistically feasible and cost efficient. In addition, this was a species specific phenomenon. HS significantly increased RDX reduction for G. sulfurreducens (Table 2.1) relative to cells alone, but
RDX reduction by G. metallireducens was not stimulated by HS at the concentrations tested. Based on past data these options include river sediment humic/fulvic extracts, leaf litter extracts, and possibly military smoke dyes that contain quinones (Rubin and
Buchanan, 1985; Nevin and Lovley, 2002a). In all cases HS and AQDS facilitated reactions were faster in the absence of bioavailable, poorly crystalline Fe(III), and
RDX was completely removed from the system in HS amended incubations; AQDS amended incubations were RDX depleted at a faster rate.
Bioavailable Fe(III) has a dual role with respect to electron shuttles and RDX.
In both cases the eventual outcome is RDX reduction and therefore degradation; however the mechanisms to reach that point differ and therefore the rates vary in the presence and absence of Fe(III). Fe(III) readily accepts electrons from reduced HS and AH 2QDS; Fe(III) reduction in this manner is very fast (Lovley et al., 1996b;
Lovley et al., 1998). As demonstrated here, Fe(III) at times competed for electrons from the reduced HS and AH 2QDS and RDX was only reduced once Fe(II) had been generated. In the absence of Fe(III) electrons are transferred directly from the reduced shuttle to RDX, with very little degradation lag time. In the presence of
Fe(III) the lag time increased and RDX degraded only once Fe(III) had been significantly reduced. Therefore the “dual role” is a slight inhibitory period followed by a stimulatory period – with respect to RDX reduction. Fe(III) reducing biomass 23 generated in these environments will reduce available HS; therefore any increase in
Fe(III) reduction eventually benefits site restoration; however, direct electron transfer from electron shuttles will be faster than from cells or Fe(II).
MNX, DNX, and TNX accumulated only transiently in either of the resting cell suspensions without Fe(III) present. In several experiments metabolites were reduced so quickly that they were not even quantified above the error range of the method detection limit. However, MNX accumulated for longer periods of time when
Fe(III) was present. MNX eventually degraded in all Fe(III) amended incubations, but MNX degradation was within several days rather than a single day, which was the case in AQDS or HS amended incubations that lacked Fe(III); DNX and TNX did not accumulate.
Bioremediation options for RDX have been limited to electron donor amendment to stimulate native microorganisms to potentially reduce RDX, most likely by direct electron transfer from cells to the nitro groups of the molecule
(Hawari et al., 2000b; Sheremata and Hawari, 2000; Zhao et al., 2003b; Bhushan et al., 2004; Zhao et al., 2004c). However, this approach can be described as moderately successful at best, with rates and extent of RDX transformation varying widely among the treated material. The data presented above describe a novel approach for RDX reduction: targeting HS and/or Fe(III) reduction by adding soluble electron shuttles to promote extracellular electron transfer from Fe(III)/HS reducing biomass to RDX
(and its nitroso intermediates.) Targeting a group of microorganisms known to be ubiquitous in their distribution the process circumvents the need for specialized cells capable of direct RDX reduction. This approach may become an efficient bioremediation strategy for subsurface environments contaminated by RDX or other cyclic nitramine compounds.
24
Figure 2.1. Structure of RDX, its nitroso metabolites (MNX, DNX, and TNX) and AQDS. RDX reduction consists of a series of two electron transfer steps for each nitro nitroso reaction
25
Figure 2.2. RDX reduction by biologically reduced AQDS (B AH 2QDS) (A) and
AH 2QDS oxidation (B) in pure phase incubations (no cells). Results are the means of triplicate analyses and bars indicate one standard deviation
26
Figure 2.3. Stoichiometric reduction of MNX (A) and DNX (B) with three different concentrations of biologically reduced AQDS (B AH 2QDS). (The DNX stock is not chemically pure; it is ca. 58% DNX, 34% MNX, and 8% TNX.). Separate incubations of MNX (C) and DNX (D) with 100 M biologically reduced AQDS (B AH 2QDS), indicating different reduction rates of DNX than later experiments of MNX or DNX (A and B). The results in panels A and B are the means of triplicate analyses, and the bars indicate one standard deviation; the preliminary experiments in panels C and D are the results of a single analysis
27
Figure 2.4. RDX reduction (A) and MNX accumulation (B) with AQDS in cell suspension incubations of G. metallireducens ; RDX reduction (C) and MNX accumulation (D) with humic substances in cell suspension incubations of G. metallireducens ; RDX reduction (E) and MNX accumulation (F) with poorly crystalline Fe(III) oxide (FeGel) in cell suspension incubations of G. metallireducens . The results are the means of triplicate analyses, and the bars indicate one standard deviation
28
Figure 2.5. RDX reduction (A), MNX accumulation (B), and Fe(II) accumulation (C) with AQDS plus poorly crystalline Fe(III) oxide (FeGel) and humic substances plus poorly crystalline Fe(III) oxide in cell suspension incubations of G. metallireducens . The results are the means of triplicate analyses, and the bars indicate one standard deviation
29 Table 2.1. Decay rates for pure phase (no cell) incubations a Pure phase (no cell) Decay rate [ mol of RDX h –1 at 100 M Temp (° C) incubation AH 2QDS or Fe(II)]
AH 2QDS alone 4.050 25 0.400 4 0.6 mM Fe(II) 0.042 25 1.2 mM Fe(II) 0.075 25 a Abiotic degradation rates of RDX by B AH 2QDS at 4°C and by biologically reduced, soluble Fe(II) are shown. Results are the means of triplicate analyses
30 Table 2.2. Decay rates for cell suspension incubations a Decay rate b Cell suspension incubation ( mol of RDX h –1 mg of cell protein –1) GS15 PCA Cells + AQDS + acetate 0.0312 0.0648 Cells + AQDS (no acetate) 0.0280 0.0565 Cells + HS + acetate 0.0079 0.0470 Cells + HS (no acetate) 0.0065 0.0391 Cells + Fe(III) + acetate 0.0002 0.0091 Cells + Fe(III) (no acetate) 0.0002 0.0083 cells + AQDS + Fe(III) + acetate 0.0131 0.0008 Cells + HS + Fe(III) + acetate 0.0009 0.0058 Cells + acetate 0.0150 0.0318 Cells (no acetate) 0.0174 ND c Aerated cells + acetate 0.0142 ND Aerated cells (no acetate) 0.0150 ND a Biodegradation rates of RDX in the resting cell suspension of G. metallireducens and G. sulfurreducens with or without extracellular electron shuttling compounds (with acetate as the sole electron donor). Results are the means of triplicate analyses. b All cell suspensions were performed at 30°C. c ND, not determined.
31 CHAPTER 3 HEXAHYDRO 1,3,5 TRINITRO 1,3,5 TRIAZINE (RDX) AND OCTAHYDRO
1,3,5,7 TETRANITRO 1,3,5,7 TETRAZOCINE (HMX) BIODEGRADATION
KINETICS AMONGST SEVERAL FE(III) REDUCING GENERA 3
INTRODUCTION
Bioremediation is one alternative for attenuating explosives contamination and several mechanisms have been proposed using pure and mixed cultures or contaminated sediment or soil (Adrian et al., 2003; Adrian and Arnett, 2004;
Sherburne et al., 2005; Thompson et al., 2005; Meyers et al., 2007). Anaerobic biodegradation in particular has been suggested as an effective strategy for RDX and
HMX because of the nitro nitroso degradation pathway (initial reductive step)
(Adrian et al., 2003) and the fact that subsurface RDX or HMX plumes can be anaerobic (Garg et al., 1991 ; Adrian and Arnett, 2004). Direct microbial RDX or
HMX reduction has been reported (Zhao et al., 2002; Zhao et al., 2003b); Boopathy and Manning (2000), and this process can be catalyzed by microorganisms that specifically transfer electrons to nitramine compounds; however, without these microorganisms, the overall biodegradation kinetics will be limited. There is still general disagreement whether microorganisms involved reduce RDX via co metabolic reactions (i.e. non energy generating reactions) or through direct respiration. It is likely some combination of both pathways, but the references above generally describe co metabolic reactions. Alternate degradation pathways are possible, but are less easily stimulated in subsurface environments (Pennington and
Brannon, 2002; Crocker et al., 2006).
3 Reproduced in part with permission from Kwon, M.J. and K.T. Finneran. Soil and Sediment Contamination . 2008, 17 : 189 – 203
32 Cyclic nitramine biodegradation mediated by Fe(III) reducing microorganisms and the extracellular electron transfer reactions catalyzed by these organisms may be another reasonable approach based on their ubiquity in subsurface environments
(Coates et al., 1998). Recently, the role of fermentative cultures in reductive transformation of explosive compounds was investigated with humic substances (HS) and anthraquinone 2,6 disulfonate (AQDS) (Borch et al., 2005; Bhushan et al., 2006).
The degradation of TNT was stimulated by adding AQDS to fermentative cultures
(Borch et al., 2005); however, TNT reduction is different than cyclic nitramine biodegradation as TNT interacts with more reduced compounds than RDX or HMX.
A fermentative bacterium, Clostridium sp . EDB2, effectively degraded RDX and
HMX (Bhushan et al., 2006). However, adding AQDS did not significantly increase the explosives reduction rates. The same group recently isolated a novel Shewanella species, but these reactions have yet to be reported with that culture (Zhao et al.,
2005). Although these data contribute to the potential strategies for explosives residues, it is unclear whether the specific fermentative organisms used within these experiments are ubiquitous in explosives contaminated environments or can be directly stimulated.
Recently, Kwon and Finneran (2006) reported that remediation strategies that manipulate Fe(III) reducers via extracellular electron shuttling are reasonable because they do not rely on organisms that must directly transfer electrons to RDX (Lovley and Phillips, 1986; Coates et al., 1995; Kashefi and Lovley, 2003). Despite the successful reduction of RDX with two Geobacter species, the rates and extent of
RDX between these organisms were different in the presence or absence of Fe(III) oxide and/or electron shuttles. In addition, since Geobacteraceae proliferate in strictly anoxic environments, RDX or HMX reduction at the fringe between aerobic and anaerobic conditions may be limited. 33 In this study, one Geobacter species ( G. metallireducens ) and three non
Geobacteraceae Fe(III) reducing genera also capable of reducing extracellular electron shuttles were investigated to identify the kinetics of RDX and HMX degradation. Anaeromyxobacter and Shewanella , gram negative facultative anaerobes, and Desulfitobacterium , gram positive strict anaerobes were used in addition to the Geobacteraceae to determine differences in RDX reduction amongst the “model” Fe(III) reducing cells and how the electron shuttling strategy will work in the presence or absence of Fe(III).
MATERIALS AND METHODS Medium and culturing conditions . The basal anaerobic medium consisted of (g/L unless specified otherwise): NaHCO 3 (2.5), NH 4Cl (0.25), NaH 2PO 4 H 2O (0.6), KCl
(0.1), modified Wolfe’s vitamin and mineral mixtures (each 10 ml/L) (Lovley et al.,
1984) and 1mL of 1mM Na 2SeO 4. Electron acceptors used with the anaerobic medium included soluble Fe(III) citrate (45 mM), poorly crystalline Fe(III) oxide (50 mmol/L), or anthraquinone 2,6 disulfonate (AQDS – 5 mM). The medium was buffered using a 30 mM bicarbonate buffer equilibrated with 80:20 (vol/vol) N 2:CO 2, as previously described (Finneran et al., 2003). Yeast extract (0.01%) was added to the culture of A. dehalogenans . Acetate (20 mM) or lactate (20 mM) was added as the sole electron donor, depending on the specific culture. Standard anaerobic and aseptic culturing techniques were used throughout (Lovley and Phillips, 1988). The aerobic medium was an LB medium that consisted of (g/L) Bacto tryptone (10.0), Bacto yeast extract (5.0), and NaCl (10.0). The medium was autoclaved prior to inoculation and cultures were transferred in 250 500 mL conical flasks. Individual experiments utilized different culture tubes or bottles, as described below.
34 Resting cell suspensions . G. metallireducens was grown anaerobically in freshwater medium with acetate as the sole electron donor and Fe(III) citrate as the sole terminal electron acceptor. A. dehalogenans was grown anaerobically in freshwater medium with acetate as the sole electron donor, Fe(III) citrate as the sole terminal electron acceptor, yeast extract (0.01%) as rich nutrients. D. chlororespirans was grown anaerobically in freshwater medium with lactate as the sole electron donor and AQDS as the sole terminal electron acceptor. S. oneidensis was grown aerobically in LB medium. One liter of each individual cell culture was harvested during logarithmic growth phase and centrifuged at 4225 g for 15 minutes to form a dense cell pellet.
Each resultant cell pellet was resuspended in 30mM bicarbonate buffer under a stream of anaerobic gas. The washed cells were centrifuged again at 4225 g for 15 minutes and the resultant biomass was resuspended in 4.0mL of bicarbonate buffer. Cells were used within 30 minutes of processing.
Experimental tubes were prepared by sparging approximately 5.0 mL of 30 mM bicarbonate buffer with anaerobic N 2:CO 2 and sealing the buffer under an anaerobic headspace. Electron acceptors or shuttles incubated with the cells including humic acids (0.25 g/L), poorly crystalline Fe(III) oxide (45 mmol/L), AQDS (5 mM), humic acid plus poorly crystalline Fe(III) oxide, and AQDS plus poorly crystalline
Fe(III) oxide. The sterilized stock solution of humic acids (10 g/L) was prepared by filtration through 0.2 m PTFE filters (humic acid did not interfere with nitramine quantitation). Cells were incubated at 30°C. An aliquot (0.3 ml) of the resting cells was added to the sealed pressure tubes to initiate each experiment.
Samples (0.3 ml) were collected periodically via anaerobic syringe and needle, and samples were filtered through sterile, 0.2 m PTFE filters (PALL Life Sciences; filters did not interfere with nitramine quantitation) prior to analyses. 0.2 ml of each sample was used to quantify RDX or HMX at each time point. 0.1 ml of samples was 35 used to quantify Fe(II) concentration in poorly crystalline Fe(III) amended incubations. No more than seven samples were taken from any incubation; therefore, final volume of each incubation was approximately 8mL at the end of the experiments
(80% volume remaining). Total cellular protein was determined by using the DC
Protein Assay (Bio RAD) and a modified Lowry protein assay (Lowry et al., 1951).
Cell protein normalized decay rates were calculated based on pseudo first order degradation coefficients. All experiments were performed in triplicate.
RESULTS
RDX degradation by Anaeromyxobacter dehalogenans strain K
A. dehalogenans reduced RDX most rapidly in the presence of AQDS (Figure
3.1A); however, RDX reduction rates by A. dehalogenans were relatively slow compared to the other cultures (Table 3.1). A. dehalogenans directly reduced RDX slowly relative to electron shuttle amended cells; more than 70% of initial RDX concentration still remained when extracellular electron shuttles were not present. A. dehalogenans did not degrade RDX effectively with poorly crystalline Fe(III) hydroxide regardless of the presence of extracellular electron shuttles (Figure 3.1B).
MNX, the one nitroso metabolite of RDX, was initially detected but quickly fell below detection limit in cells plus HS incubations and cells only incubations. Nitroso metabolites did not accumulate in AQDS and/or Fe(III) amended incubations (Table
3.2).
RDX degradation by Shewanella oneidensis strain MR1
RDX was completely reduced within 3 h in S. oneidensis suspensions containing AQDS (Figure 3.2A). HS amended RDX reduction rates were relatively slow compared to the other cultures. S. oneidensis also used RDX directly as an 36 electron acceptor. RDX degradation rates were similar regardless of the presence or absence of lactate (Figure 3.2A and 3.2B). RDX in poorly crystalline Fe(III) amended incubations was reduced to 30 M in 65 h; however, AQDS and HS increased RDX reduction when Fe(III) was present (Figure 3.2B). All incubations except the AQDS amended suspensions accumulated nitroso metabolites at concentrations higher than other cultures (Table 3.2).
RDX degradation by Desulfitobacterium chlororespirans strain Co23
D. chlororespirans completely reduced RDX within 4 h and 45 h in suspensions that contained AQDS and HS, respectively (Figure 3.3A). RDX was not completely reduced in any other treatment. Cells alone (with lactate as the electron donor) reduced RDX to 20 M in 43 h, but further reduction was limited. D. chlororespirans did not reduce RDX in the absence of electron donor (lactate). RDX was reduced in the presence of Fe(III); however, this rate was increased by adding
AQDS or HS to Fe(III) incubations (Figure 3.3B). Nitroso metabolites accumulated in the presence of HS and/or Fe(III). Nitroso metabolites did not accumulate in AQDS amended incubations without Fe(III) or within the cells only incubations (Table 3.2).
HMX degradation by G. metallireducens
The initial concentration of HMX in the resting cell suspensions of G. metallireducens was ca. 7 M, due to the low solubility of HMX in water.
Extracellular electron shuttles stimulated HMX reduction; 7 M HMX was reduced to approximately 0.5 M in 35 h in AQDS or HS amended incubations (data not shown). Cells only incubations reduced HMX to 4 M in the same time period.
Resting cell suspensions with only poorly crystalline Fe(III) oxide reduced HMX to
4.5 M in 35 h, while adding AQDS to Fe(III) increased HMX reduction rates (Table 37 3.1). Nitroso metabolites of HMX accumulated in the presence of HS and/or Fe(III), while their accumulation in AQDS only amended incubations and cells only incubations was insignificant (Table 3.2).
DISCUSSION These results demonstrate that four genera of Fe(III) reducing Bacteria transform RDX (and one at least HMX) and that extracellular electron shuttles increase the rate and extent of RDX and HMX biodegradation catalyzed by Fe(III) reducing microorganisms. Cyclic nitramine biodegradation (directly and via extracellular electron transfer) is ubiquitous among the “model” Fe(III) reducing
Bacteria (i.e. Fe(III) reducing microorganisms that dominate anoxic environments both in cell density and phylogenetic distribution). This suggests that cyclic nitramines may not persist in subsurface environments dominated by Fe(III) reduction, particularly when extracellular electron shuttling compounds are present.
Although it is possible to stimulate the reduction of explosives with the addition of organic carbon alone, the reported degradation rates of cyclic nitramines under anaerobic conditions have varied (Hawari et al., 2000b; Zhao et al., 2002; Wani and
Davis, 2003; Wani and Davis, 2006). Electron shuttling can promote nitramine reduction by accepting electrons in microbial respiration and abiotically transferring the electrons to RDX or HMX at rates significantly faster than those previously reported. The electron shuttles are re oxidized and available again for microbial respiration; in this manner the shuttles are catalytic and only a small concentration would be needed to promote these reactions in bioremediation applications.
RDX reduction rates were slightly different between two previously tested strains (Kwon and Finneran, 2006); G. sulfurreducens degraded RDX 2 or 6 times faster than G. metallireducens when amended with AQDS or HS, respectively. HS
38 had a more marked effect on G. sulfurreducens ; however, this may be due to the fact that cells alone of G. metallireducens more efficiently reduced RDX. Reduction rates among the species tested above were also different.
Several of the genera reduced RDX efficiently without an added electron donor. This is likely due to a phenomenon called endogenous respiration, in which the cells utilize decayed cellular biomass as a substrate for respiratory processes
(which is common for cell suspensions) (Fredrickson et al., 2000a). The organisms tested do not form storage molecules under the conditions tested. RDX reduction was also not attributed to non specific reduction via reduced membrane cytochromes, as air oxidized cells did not behave differently when incubated in the cell suspension.
D. chlororespirans was the exception, and it required an electron donor to promote these reactions. Further investigation is necessary to identify the reason that the sole gram positive member of the experiments did not follow the same endogenous respiration trend.
Degradation of RDX/HMX by Fe(III)-reducing microorganisms
Anaeromyxobacter species are a newly defined group of facultative Fe(III) and electron shuttle reducing Bacteria. The full extent of their capacity to reduce extracellular quinones and HS has not been characterized, but those tested respire and grow via AQDS reduction (data not shown). Cells directly reduced RDX (Table 3.1).
Although AQDS and HS increased the rate and extent of RDX reduction, the degradation rates of RDX by Anaeromyxobacter species were relatively slow compared to other species tested in this study. Given that they are facultative anaerobes, they may also be relevant in subsurface environments where dominant electron accepting processes are shifting, such as the boundary between aerobic and anaerobic zones. 39 Shewanella represent one of the model Fe(III) and electron shuttle reducing genera. In addition, several reports suggest RDX reducing enrichments from marine sediment were dominated by Shewanella species, and an isolate from the enrichment was within the genus Shewanella (Zhao et al., 2004c; Zhao et al., 2005). Shewanella are reported to produce soluble electron transfer molecules which can reduce electron acceptors outside of the cell membrane (Hernandez and Newman, 2001). Perhaps this metabolic capacity makes Shewanella particularly adept with cyclic nitramines, which are reduced quickly by extracellular electron shuttling compounds. Also, like
Anaeromyxobacter , the Shewanella are facultative anaerobes, which suggests that they may dominate RDX (or other nitramines) reduction at the fringe between aerobic and anaerobic conditions.
D. chlororespirans is a gram positive, low G+C content, spore forming bacterium (Sanford et al., 1996). It is reported to reduce HS but to date has not been reported to reduce Fe(III). Other members of the genus including D. metallireducens and D. hafniense can reduce Fe(III) as well as HS (Beaudet et al., 1998; Niggemyer et al., 2001; Finneran et al., 2002; Rhee et al., 2003; Shelobolina et al., 2003). D. chlororespirans can ferment pyruvate, and it grows via respiration with several electron acceptors (Sanford et al., 1996). Although D. chlororespirans is the only gram positive among organisms tested, it also directly reduced RDX to some extent suggesting cell membrane/cell wall differences did not inhibit direct RDX reduction.
AQDS and HS mediated RDX reduction was very fast (Table 1). AQDS and HS also promoted RDX reduction when Fe(III) oxide was present. Given that D. chlororespirans cannot directly reduce Fe(III) oxide electron transfer from the reduced electron shuttles was probably split between Fe(III) and RDX, which accounts for the differences in the rate and extent of RDX degradation when Fe(III) is present. 40 G. metallireducens stimulated HMX reduction with extracellular electron shuttles (Table 3.1). HMX was reduced much more slowly than RDX, which is likely due to the presence of an additional nitro group and a larger molecular structure.
However, the trends of HMX reduction were similar to those of RDX reduction with respect to electron shuttles and Fe(III). These data suggest that the direct or indirect electron transfer via extracellular electron shuttles can increase the reduction rates of other explosives containing nitro groups in their structure.
Nitroso metabolites (MNX, DNX, and TNX) of RDX accumulated to different extents among the five microorganisms tested and among different amendments
(Table 3.2). In the resting cell suspensions of G. metallireducens , nitroso metabolites of HMX also accumulated to different extents among different treatments (Table 3.2).
The accumulation of nitroso metabolites of RDX (or HMX) in AQDS amended incubations was not significant or was not detected regardless of different microbial species. Nitroso metabolites accumulated to a greater extent in HS and/or poorly crystalline Fe(III) amended incubations. Other than the AQDS/donor incubations, the decay rates of RDX in two facultative microorganisms ( Anaeromyxobacter and
Shewanella ) were similar. However, more nitroso metabolites accumulated in
Shewanella suspensions than Anaeromyxobacter suspensions.
Competition for electrons between RDX and Fe(III)
Fe(III) oxides influenced RDX reduction in the presence or absence of electron shuttles. Reduced shuttles transfer electrons quickly to Fe(III) (Kwon and
Finneran, 2006); however, these and previous data indicate that electron transfer to
RDX is also very fast (Kwon and Finneran, 2006). The data suggest that there are two possible pathways for RDX reduction: 1) reduced electron shuttles transfer electrons to Fe(III) and RDX is subsequently reduced by surface bound Fe(II) (which 41 has been reported by others (Gregory et al., 2004; Williams et al., 2005), or 2) reduced electron shuttles are simultaneously transferring electrons to both Fe(III) and
RDX. Although there may be some combination of both, the second pathway is likely correct based on reaction kinetics and metabolite accumulation patterns. The exact mechanism of electron shuttle mediated RDX/HMX reduction and the dynamic between Fe(III) and the cyclic nitramines (with electron shuttles) is being investigated in further detail.
The nitroso metabolites MNX, DNX, and TNX also accumulated to higher concentrations in systems where Fe(II) was the primary reductant (Gregory et al.,
2004; Williams et al., 2005; Kwon and Finneran, 2006). The sequential reduction of
HMX to the nitroso metabolites (1 NO HMX, octahydro 1,3 dinitroso 5,7 dinitro
1,3,5,7 tetrazocine ( 2 NO HMX), occasionally octahydro 1,3,5 trinitroso 7 nitro
1,3,5,7 tetrazocine ( 3 NO HMX), and octahydro 1,3,5,7 tetranitroso 1,3,5,7 tetrazocine ( 4 NO HMX)) by zero valent iron under anoxic conditions has also been reported (Monteil Rivera et al., 2005). Nitroso metabolite accumulation was low or undetectable with AQDS or HS, which suggests faster reduction through the nitroso series to ring cleavage products.
This dynamic between Fe(III) and RDX has implications for the in situ strategy; however, the results will likely be similar despite the electron transfer pathway. Both Fe(II) and reduced electron shuttles reduce RDX. The kinetic differences do not preclude using these strategy in sediments with high bioavailable
Fe(III). The data clearly indicate that adding electron shuttles increases RDX reduction when Fe(III) is present. Although the number of possible reactions increases in heterogeneous environments, the preponderance of data demonstrates that stimulating Fe(III) reduction increases RDX and HMX transformation.
42 60
50
40 M) 30 RDX ( RDX cell+AQDS+Ac 20 cell+AQDS-Ac cell+humics+Ac A cell+humics-Ac 10 cell+Ac cell-Ac no cell+Ac 0 60
50
40 M) 30 RDX ( RDX 20 cell+Fe(III)+Ac cell+Fe(III)-Ac B 10 cell+AQDS+Fe(III)+Ac cell+humics+Fe(III)+Ac
0 0 20 40 60 80
time (h) Figure 3.1. RDX reduction in resting cell suspensions of Anaeromyxobacter dehalogenans strain K with or without extracellular electron shuttling compounds (acetate (Ac) as the sole electron donor) in the absence (A) or presence (B) of poorly crystalline Fe(III); Results are the means of triplicate analyses and bars indicate one standard deviation.
43
40 cell+AQDS+Lc cell+AQDS-Lc cell+humics+Lc 30 cell+humics-Lc cell+Lc cell-Lc
M) no cell+Lc 20 RDX ( RDX
10 A
0
40
30 M)
20 RDX ( RDX
cell+Fe(III)+Lc 10 cell+Fe(III)-Lc B cell+AQDS+Fe(III)+Lc cell+humics+Fe(III)+Lc 0 0 10 20 30 40 50 60 70
time (h) Figure 3.2. RDX reduction in resting cell suspensions of Shewanella oneidensis strain MR1 with or without extracellular electron shuttling compounds (lactate (Lc) as the sole electron donor) in the absence (A) or presence (B) of poorly crystalline Fe(III); Results are the means of triplicate analyses and bars indicate one standard deviation.
44
40
30 M)
20 cell+AQDS+Lc cell+AQDS-Lc RDX ( RDX cell+humics+Lc cell+humics-Lc 10 cell+Lc cell-Lc A no cell+Lc
0
40
30 M)
20 cell+Fe(III)+Lc
RDX ( RDX cell+Fe(III)-Lc cell+AQDS+Fe(III)+Lc 10 cell+humics+Fe(III)+Lc B
0 0 10 20 30 40
time (h) Figure 3.3. RDX reduction in resting cell suspensions of Desulfitobacterium chlororespirans strain Co23 with or without extracellular electron shuttling compounds (lactate (Lc) as the sole electron donor) in the absence (A) or presence (B) of poorly crystalline Fe(III); Results are the means of triplicate analyses and bars indicate one standard deviation.
45 s n a . e d e m t n . i e u G h o m t f r h o e t with or wi ND, not dete b (2006). D. chlororespirans chlororespirans D. , and S. oneidensis S. s of RDX or HMX in the resting cell suspension or lactate as the sole electron donor). Results ar , A. dehalogenans A. , The results were reproduced from Kwon and Finneran a G. sulfurreducens G. ,
Table 3.1. Cell protein normalized degradation rate metallireducens extracellular electron shuttling compounds (acetate of triplicate analyses.
46 . s d e n s r l a y a l l l , s e a u e c l 8 X s n l
7 . e g a e N G c h e n e D i t r a t f t , e n r a s o t e c e X w x r i n r l e s a N o p e i t n p i s h M a r u e t n r t ) o i e h f n h t p p i t o M n with or wi D. chlorores D. 4NO HMX. micromolar ( c , and D. chlororespirans D. S. oneidensis S. , NT, not detected. , and b (2006). bolites of RDX or HMX in the resting cell sus ch nitroso metabolites. Results are the means A. dehalogenans A. S. oneidensis S. , , the sole electron donor). Total incubation times 17, 71, 65, and 43 h, respectively. The numbers i A. dehalogenans A. , G. sulfurreducens G. , M) 1NO HMX and area units for 2NO HMX, 3NO HMX, and G. sulfurreducens G. , G. metallireducens G. micromolar ( d The results were reproduced from Kwon and Finneran Table 3.2. Maximum accumulation of the nitroso meta metallireducens electron shuttling compounds (acetate or lactate as suspensions of (except cells plus Fe(III) plus donor incubations), indicate the time (h) at maximum accumulation of ea a TNX.
47 CHAPTER 4 BIOTRANSFORMATION PRODUCTS AND MINERALIZATION
POTENTIAL FOR HEXAHYDRO 1,3,5 TRINITRO 1,3,5 TRIAZINE (RDX)
IN ABIOTIC VERSUS BIOLOGICAL DEGRADATION PATHWAYS WITH
ANTHRAQUINONE 2,6 DISULFONATE (AQDS) AND GEOBACTER
METALLIREDUCENS 4
INTRODUCTION
Crocker et al reviewed the pathways by which RDX and its nitroso metabolites can be degraded (Crocker et al., 2006). Hawari et al (2000) (Hawari et al., 2000b) demonstrated that municipal anaerobic sludge can degrade RDX to hexahydro 1 nitroso 3,5 dinitro 1,3,5 triazine (MNX) and hexahydro 1,3 dinitroso 5 nitro 1,3,5 triazine (DNX), or to ring cleavage products. Phanerochaete. chrysosporium transformed octahydro 1,3,5,7 tetranitro 1,3,5,7 tetrazocine (HMX) to octahydro 1 nitroso 3,5,7 trinitro 1,3,5,7 tetrazocine (1 NO HMX) plus ring cleavage products in nitrogen limited medium (Fournier et al., 2004b). A facultatively anaerobic bacterium isolated from anoxic sludge, Klebsiella pneumoniae strain SCZ 1, degraded RDX by an initial denitration reaction followed by ring cleavage and decomposition in water (Zhao et al., 2002); RDX transformation to nitroso metabolites was minor. A homoacetogen, Acetobacterium paludosum , degraded RDX fastest under autotrophic (H 2 fed) conditions when nitrogen sources
(e.g. ammonium) were absent. MNX, DNX, and hexahydro 1,3,5 trinitroso 1,3,5 triazine (TNX) were not detected. N 2O accumulated to 64% of the total nitrogen, but
RDX was not mineralized to CO 2 (Sherburne et al., 2005).
4 Reproduced in part with permission from Kwon, M.J. and K.T. Finneran. Biodegradation . 2008, 19 : 705 – 715
48 Cyclic nitramine biodegradation mediated by extracellular electron shuttling
(primarily anthraquinone 2,6 disulfonate (AQDS)) and the microorganisms that catalyze these reactions may be another reasonable approach based on the ubiquity of these microbes in subsurface environments (Coates et al., 1998). Recently, the role of fermentative cultures in reductive transformation of explosive compounds was investigated with AQDS (Borch et al., 2005; Bhushan et al., 2006). TNT degradation was stimulated by adding AQDS to fermentative cultures (Borch et al., 2005). A fermentative bacterium, Clostridium sp . EDB2, effectively degraded RDX (Bhushan et al., 2006) with AQDS present. Adding AQDS did not significantly increase the
RDX reduction rates, but the low pH conditions generated by fermentation may have limited anthrahydroquinone 2,6 disulfonate (AH 2QDS) mediated RDX reduction.
Bradley and Dinicola suggested that Mn(IV) reduction was a dominant pathway for
RDX biodegradation, and Mn(IV) reducing microorganisms are closely related to the known electron shuttle reducers (Bradley and Dinicola, 2005).
One past report (Schwarzenbach et al., 1990) and more recent studies
(Uchimiya and Stone, 2006) demonstrate that reactivity of hydroquinone compounds is pH dependent. Schwarzenbach et al (1990) demonstrated that the rate of reduction of nitroaromatic compounds by Juglone and lawsone increases with pH, and
Uchimiya and Stone (2006) reported that oxidation of 1,2 naphthoquinone 4 sulfonate by 2,6 dimethylhydroquinone was pH dependent. In the current study, the effect of pH on RDX bioremediation mediated by extracellular electron shuttling compounds was investigated because contaminated subsurface environments with different pH values will influence overall degradation.
Our previous study demonstrated RDX bioremediation by extracellular electron shuttling compounds with two different bacterial species (Kwon and
Finneran, 2006). However, it did not address the ring cleavage products in 49 extracellular electron shuttle amended systems. Specifically, what are the reaction products with the abiotic pathway versus the biological pathway? When both pathways are available (abiotic plus biotic, the operationally defined mixed pathway)
what are the dominant products? These data are critical to developing this bioremediation strategy in terms of monitoring and extent of transformation. The strictly abiotic, strictly biological, and mixed pathways were investigated separately at three circumneutral pH values to quantify the ring cleavage products of RDX degradation generated by each. Geobacter metallireducens was used as the RDX and electron shuttle reducing model microorganism and AQDS was used as the sole electron shuttle.
MATERIALS AND METHODS Culturing conditions . Geobacter metallireducens strain GS 15 (ATCC 53774) was maintained using Fe(III) Citrate or AQDS media described below. The basal anoxic
1 medium consisted of (g l unless specified otherwise): NaHCO 3 (2.5), NH 4Cl (0.25),
NaH 2PO 4 H 2O (0.6), KCl (0.1), modified Wolfe’s vitamin and mineral mixtures (each
1 10 ml l ) and 1 mL of 1 mM Na 2SeO 4 (Lovley et al., 1993). Electron acceptors used with the anoxic medium were soluble Fe(III) citrate (45mM) or anthraquinone 2,6 disulfonate (AQDS – 5 mM). The medium was buffered using a 30mM bicarbonate buffer equilibrated with 80:20 (vol/vol) N 2:CO 2, as previously described (Finneran et al., 2003). Acetate (20 mM) was added as the sole electron donor. Standard anoxic and aseptic culturing techniques were used throughout (Lovley and Phillips, 1988).
Pure phase incubations . Biologically reduced AQDS (B AH 2QDS) was prepared by incubating Geobacter metallireducens in AQDS medium (5 mM) . The B AH 2QDS was filtered through a 0.2 m sterilized PTFE filter into a pre sterilized, anoxic serum bottle to remove cells. Experimental tubes were prepared by sparging approximately
50 7.0ml of 30mM bicarbonate buffer with anoxic N 2:CO 2 (= 100 : 0 for pH 9.2, 99.9 :
0.1 for pH 8.2, 99.6 : 0.4 for pH 7.9, 96.3 : 3.7 for pH 7.3, 80 : 20 for pH 6.8 and 40 :
60 for pH 6.2). The buffer was sealed under an anoxic headspace of the same mixture.
The concentration of reduced AQDS tested was 100 M for the all incubations with
RDX (33 M – 39 M). Incubations were performed at 30°C. Samples were collected periodically via anoxic syringe and needle; samples were filtered through sterile, 0.2 m PTFE filters prior to analyses (PALL Life Sciences; filters did not interfere with nitramine or metabolite quantitation).
Resting cell suspensions . G. metallireducens was grown anaerobically in freshwater medium with acetate as the sole electron donor and Fe(III) citrate as the sole terminal electron acceptor. One liter of cell culture was harvested during logarithmic growth phase and centrifuged at 5250 x g for 15 minutes to form a dense cell pellet. The cell pellet was resuspended in 30mM bicarbonate buffer under a stream of anoxic gas.
The washed cells were centrifuged again at 5250 x g for 15 minutes and the resultant biomass was resuspended in 4.0mL of bicarbonate buffer. Cells were used within 30 minutes of processing. There was no residual Fe(II) present in the suspension that could act as an abiotic reductant.
Experimental tubes were prepared by sparging approximately 5.0ml of 30mM bicarbonate buffer with anoxic N 2:CO 2 as described above for pH 9.2, 7.9, and 6.8 and sealed under an anoxic headspace. Acetate (20mM) was amended as the sole electron donor. AQDS (100 M) was the sole electron shuttle added to the cell suspensions. Cells were incubated at 30°C. An aliquot (0.3ml) of the resting cells
(0.27 ± 0.02 mg protein/ml) was added to the sealed pressure tubes to initiate each experiment. The final volume in each tube was 10.0mL. The starting RDX concentration was approximately 60 M.
51 Samples (0.5 ml) were collected periodically via anoxic syringe and needle, and samples were filtered through sterile, 0.2 m PTFE filters prior to analyses. To minimize sampling volume, glass inserts (250uL Glass LVI Flat Bottom; Laboratory
Supply Distributors, NJ) were used in the autosampler vials. 0.05 ml of sample was used to quantify RDX and its nitroso metabolites (MNX, DNX, and TNX) at each time point. 0.05, 0.1, 0.1, and 0.1 ml of samples were used to quantify
methylenedinitramine (MEDINA), formaldehyde (HCHO), nitrite (NO 2 ), ammonium
+ (NH 4 ), respectively. No more than five samples were taken from any incubation; therefore, the final volume of each incubation was approximately 7.5 mL at the end of the experiments (75% volume remaining). Nitrous oxide and methanol were monitored by headspace analysis. All experiments were performed in triplicate.
[14 C] RDX study with resting cell suspension. Cell suspension incubations (10 mL)
14 14 were prepared for CO 2 and CH4 analysis. pH was controlled with bicarbonate buffer by N 2:CO 2 gas bubbling as mentioned above. Acetate (0.5 mM) was amended as the sole electron donor. AQDS (100 M) was the only electron acceptor incubated with the cells; the control was cells alone (no electron acceptor). U [14 C] RDX was
14 amended at the concentration for a final radioactivity of 18,000 dpm/ml. CO 2 and
14 14 CH4 were monitored by analysis of headspace samples (1ml). H CO 3 was used to
14 establish CO 2 partitioning between the liquid and gas phase, and was factored in to
14 the final mineralization; the partition coefficients (total dpm H CO 3 recovered as
14 CO 2/total dpm added) were 0.005, 0.027, and 0.135 at pH 9.2, 7.9, and 6.8,
14 respectively. The experimental tubes at pH 9.2 were acidified to convert H CO 3 to
14 CO 2 at the end of the experiment (acidified partition coefficient = 0.340). All experiments were performed in triplicate.
52 RESULTS pH effects on RDX reduction and AH 2QDS oxidation
As pH increased with AH 2QDS fixed at 100 M the rates of RDX reduction
st 1 increased (Figure 4.1). Pseudo 1 order rate constants (k obs ; h ) at pH 9.2, 8.2, 7.9,
7.3, 6.8, and 6.2 were 0.428±0.004, 0.168±0.012, 0.118±0.002, 0.054±0.006,
0.019±0.001, and 0.007±0.001, respectively. Comparing the fractional concentration of the monoprotonated hydroquinone ( α1) versus k obs in a log log plot demonstrates that the rate is proportional to α1 (the fraction of the monoprotonated form) (Figure
4.2). α1 was calculated as:
α1 = [monoprotonated reduced AQDS]/C total ,
+ + [monoprotonated reduced AQDS] = C total /([H ] / K a1 + 1 + K a2 /[H ]),
Ctotal = [diprotonated reduced AQDS] + [monoprotonated reduced AQDS] +
[unprotonated reduced AQDS] = 100 M,
8.1 10.5 Where K a1 = 10 and K a2 = 10
Transformation products in the abiotic, biological, and mixed abiotic-biological pathways
RDX reduction increased in strictly abiotic suspensions with AH 2QDS
(100 M) as the pH increased from 6.8 to 9.2 (Figure 4.3A). HCHO and MEDINA also increased, while MNX production was negligible (Figure 4.3B and 4.3C and
Table 4.1). The inorganic nitrogen compounds generated were nitrite, nitrous oxide,
14 14 and ammonium (Figure 4.4A, 4.4B, and 4.4C). CO 2 or CH 4 were not produced from 14 C labeled RDX in any of the abiotic suspensions (Table 4.1).
In contrast to RDX reduction by the abiotic pathway, cells alone degraded
RDX faster at pH 6.8 and 7.9 than at pH 9.2 (Figure 4.3D). HCHO was the primary 53 carbon metabolite and MNX did not accumulate; however, MEDINA was less significant relative to the strictly abiotic pathway and was depleted at pH 6.8 (Figure
4.3E and 4.3F). Methanol (CH 3OH) was a limited product in the carbon mass balance at pH 9.2 (Table 4.1). Inorganic nitrogen was predominately ammonium at all pH tested (Figure 4.4E). More nitrous oxide did accumulate at lower pH (Figure 4.4F).
14 14 CH 4 was not produced from C labeled RDX in any of the cell suspensions (Table
14 4.1), but CO 2 was produced.
Cells + AQDS (the operationally defined “mixed” pathway) degraded RDX quickly and the rates and extent of RDX reduction were similar at all pH tested
(Figure 4.3G). HCHO and MEDINA were concurrently produced (Figure 4.3H and
4.3I); however, MEDINA started to decline at different times for each pH and continued to decline through the final sampling points (an additional data point at hour 66 was not shown) (Figure 4.3I). Methanol was also produced at about the same
14 percentage of the mass balance as the cells alone suspension. CH 4 was not
14 14 produced from C labeled RDX in any of the cells + AQDS suspensions. CO 2 was produced from 14 C labeled RDX in all cells + AQDS suspensions (Table 4.1). Nitrite was produced and disappeared quickly at pH 9.2, while nitrite was not detected at pH
7.9 and 6.8 (Figure 4.4G); ammonium was the dominant inorganic nitrogen product
(Figure 4.4H).
DISCUSSION RDX degrades to a variety of intermediates under anaerobic conditions, influenced by the microbial community present and the geochemical conditions
(Crocker et al., 2006). The critical issue for RDX remediation is how fast the ring cleavage metabolites are produced, and whether they are amenable to further degradation or ultimately mineralization. Formaldehyde and methanol are two
54 desirable carbon intermediates, as neither will persist in subsurface systems. The inorganic nitrogen intermediates are labile, but each can serve as an indicator of the extent of ring cleavage. Our results demonstrate that electron shuttling compounds
(whether in strictly abiotic reactions or in mixed biological abiotic systems) accelerated formaldehyde production and mineralization relative to pathways devoid of electron shuttling compounds. While this study targets one specific microbial community (extracellular electron shuttle reducers), it is a reasonable strategy for
RDX bioremediation and it demonstrates that the rate and extent of ring cleavage and mineralization can be influenced.
pH influence on RDX reduction in the abiotic, biological, and mixed abiotic biological pathways
Increasing pH from 6.8 to 9.2 in the abiotic pathways increased the rates and extent of RDX reduction and metabolite formation. Increasing pH in the strictly biological pathway decreased RDX reduction rates and metabolite distribution.
However, in the mixed abiotic biological pathway the rates and extent were similar regardless of pH, which suggests an optimal pathway in the presence of cells and shuttles. The pH range was selected because it is within the normal tolerance range of
G. metallireducens (and other AQDS reducing cells) and was below the pH where alkaline hydrolysis of RDX has been reported (pH ≥ 10.5) (Hwang et al., 2006).
While G. metallireducens is metabolically active from pH 6.8 9.2 its activity is sub optimal at the higher pH values (pH ≥ 8.0).
The pH dependence of RDX reduction by AH 2QDS (abiotic pathway) is consistent with the findings by Schwarzenbach et al (1990) where they reported pH dependence on the rate of nitroaromatic compound reduction in the presence of
1 juglone and lawsone; as pH increased from 6.1 to 7.8, k obs (s ) increased. Uchimiya 55 and Stone (2006) also reported that monoprotonated dihydroxybenzene molecules have greater reactivity than diprotonated molecules and that forward rate constants of
1,2 naphthoquinone 4 sulfonate and 1,4 naphthoquinone 2 sulfonate reduction by
2,6 dimethylhydroquinone increase as pH increases (> pH 4.0, > pH 5.0, respectively)
(Uchimiya and Stone, 2006). They postulated that the relative proportion of the monoprotonated hydroquinone is greater than the diprotonated hydroquinone, which increases quinone mediated reduction rates. In the current study a fixed concentration of AH 2QDS (100 M) reduced RDX at a higher rate and extent as pH increased.
Between pH 8.5 and 9.2 approximately 90% of the anthrahydroquinone is in the monoprotonated ( α1) form (data not shown). Plotting log α1 versus log kobs (for RDX reduction) demonstrates a linear relationship between the RDX reduction rate and the proportion of the monoprotonated hydroquinone (Figure 4.2). Although some electrons must be transferred from the un protonated ( α2) form of the hydroquinone, this plot assumes that α1 >> α2. At values higher than pH 10, alkaline hydrolysis of
RDX is a dominant pathway; therefore, the un protonated form of the hydroquinone is difficult to test.
Transformation products in the abiotic, biological, and mixed abiotic-biological pathways
Formaldehyde was the dominant carbon product in electron shuttle amended suspensions (abiotic or mixed). Formaldehyde production in the absence of electron shuttling compounds was lower, particularly at the higher pH values tested (pH 7.9 and 9.2). This is critical because formaldehyde is one of the lowest molecular mass metabolites during RDX breakdown, and is a precursor to mineralization (Hawari et al., 2000a; Sherburne et al., 2005). MEDINA was also produced, but declined
14 14 slightly in the presence of cells (Figure 4.3). C RDX was mineralized to CO 2 to 56 some extent in all cells amended incubations. The mass balance for cell containing suspensions was not closed, primarily at low pH; it is possible that there were unidentified intermediates. Biosorption was considered but little radioactivity (less than 1.0%) was recovered from cell pellets obtained from each suspension in the
[14 C] RDX experiments (data not shown). Figure 4.5 presents a proposed pathway for dominant products and metabolites for the abiotic and mixed pathways involving electron shuttles.
Ammonium was the dominant inorganic nitrogen product in cell containing suspensions at all pH tested, and AQDS increased the extent of ammonium produced
(especially at pH 9.2), most likely by stimulating the rate at which nitrite was produced compared to cells alone. Nitrite was a significant end product in the strictly abiotic pathway with AH 2QDS at all pH tested. Nitrite reduction and ammonium production in cell suspensions are not surprising since Geobacter can reduce nitrate and nitrite to ammonium (Lovley, 2000a); nitrite reductase activity in its membrane is well established (Martínez Murillo et al., 1999). Ammonium was produced to a much lower extent in the AH 2QDS alone suspensions. Nitrous oxide was produced in all suspensions. Nitrous oxide production from RDX, MNX or MEDINA by abiotic and biological pathways has been reported previously (Bhushan et al., 2002; Fournier et al., 2002; Zhao et al., 2002).
MEDINA and 4 nitro 2,4 diazabutanal (NDAB) have also been reported as ring cleavage intermediates (Hawari et al., 2000b; Zhao et al., 2003b). However, formaldehyde accounted for a greater proportion of carbon than MEDINA in all suspensions suggesting that MEDINA will be less significant than HCHO (Table 4.1) under electron shuttle reducing conditions. MEDINA may spontaneously decompose in water to nitrous oxide and formaldehyde or it can be also transformed directly by cells (Halasz et al., 2002), which will be further evaluated for G. metallireducens . 57
Figure 4.1. Abiotic hexahydro 1,3,5 trinitro 1,3,5 triazne (RDX) reduction by
AH 2QDS in cell free incubations at pH 6.2, 6.8, 7.3, 7.9, 8.2, and 9.2. Results are the means of triplicate analyses and bars indicate one standard deviation.
58
Figure 4.2. Correlation between the fractional concentration of monoprotonated reduced AQDS ( ααα1) and pseudo first order degradation coefficients at pH 6.2, 6.8, 7.3, 7.9, 8.2, and 9.2 (pH increases from left to right along the x axis). The solid line is the linear regression analysis. The dashed lines indicate 95% confidence intervals. Results are the means of triplicate analyses and bars indicate one standard deviation.
59
Figure 4.3. RDX reduction and HCHO/MEDINA production by abiotic (A, B, and C), biological (D, E, and F), mixed abiotic biological (G, H, and I) pathways at pH 6.8, 7.9, and 9.2. For biological or mixed abiotic biological pathways RDX reduction was tested in resting cell suspension of G. metallireducens with (G, H, I) or without (D, E, F) AQDS. Acetate was the sole electron donor. Results are the means of triplicate analyses and bars indicate one standard deviation. Note that 40 M RDX was the starting concentration in the strictly abiotic series, and that 60 M RDX was the starting concentration in both series with cells.
60
Figure 4.4. Production of nitrite, ammonium, and nitrous oxide from RDX by abiotic (A, B, and C), biological (D, E, and F), mixed abiotic biological (G, H, and I) pathways at pH 6.8, 7.9, and 9.2. For biological or mixed abiotic biological pathways RDX reduction was tested in the resting cell suspension of G. metallireducens with (G, H, I) or without (D, E, F) AQDS. Acetate was the sole electron donor. Results are the means of triplicate analyses and bars indicate one standard deviation. Note that 40 M RDX was the starting concentration in the strictly abiotic series, and that 60 M RDX was the starting concentration in both series with cells.
61
Figure 4.5. Probable RDX degradation routes by mixed abiotic biological pathways based on products identified and reported degradation pathways (references listed on figure) in the presence of AQDS and G. metallireducens . Compounds in square brackets were not determined; [C] represents unidentified carbon intermediates. A = abiotic pathway(s), B = biological pathway(s). The >, <, and ≈≈≈ indicate whether the products were more significant in the abiotic or biological pathway (based on the mass balance).
62 r e o t
. a F c G i . l f 2 . p o d i 9 r n n t a d o f i n n s o a o n s i , e t 9 n n donor. Results are the mea reduction was tested in the resting cell susp the final sampling point during RDX reduc ixed abiotic biological pathways at pH 6.8, 7. n/a = not applicable a with or without AQDS. Acetate was the sole electro
Table 4.1. Nitrogen and carbon mass balance (%) at metabolite production by abiotic, biological, and m biological or mixed abiotic biological pathways RDX metallireducens analyses with standard deviation.
63 CHAPTER 5
HEXAHYDRO 1,3,5 TRINITRO 1,3,5 TRIAZINE (RDX) REDUCTION IS
CONCURRENTLY MEDIATED BY DIRECT ELECTRON TRANSFER
FROM HYDROQUINONES AND THE RESULTING BIOGENIC FE(II)
FORMED DURING ELECTRON SHUTTLE AMENDED
BIODEGRADATION 5
INTRODUCTION
Bioremediation is one suggested technology for RDX and other explosive contaminants. Previous studies have investigated electron shuttling via fermentative microorganisms (Borch et al., 2005; Bhushan et al., 2006). The data demonstrated electron shuttle mediated RDX transformation and contributed to the potential strategies for explosives residues. However, it is unclear whether the specific fermentative organisms utilized within these experiments are competitive in Fe(III) reducing environments contaminated with the explosive compounds. Fermentative microorganisms do not gain energy and grow via reactions with electron shuttles or iron, and may not directly participate in electron shuttling reactions in situ. Therefore, we have specifically investigated extracellular electron shuttling via Fe(III) reducing microorganisms as one possibility to increase the rate and extent of RDX transformation (Kwon and Finneran, 2006, 2008a, 2008b).
Once the initial reductive step is complete, whether it is via nitro group reduction or denitration, RDX is much less stable and is prone to spontaneous decomposition in aqueous systems (McCormick et al., 1981; Hawari et al., 2000a;
Crocker et al., 2006). Electron shuttling mediated remediation proceeds via initial 5 Reproduced in part with permission from Kwon, M.J. and K.T. Finneran. Environmental Engineering Science , “accepted”
64 nitro group reduction, and the downstream reactions are faster than spontaneous decomposition reactions (Kwon and Finneran, 2008a). Recent reports with green rusts (Larese Casanova and Scherer, 2008), Fe(II) ligand complexes (Kim and
Strathmann, 2007), and zero valent iron (Naja et al., 2008) demonstrate that electron transfer from reduced iron stimulates similar reactive rates.
All previous studies raise a question about electron shuttle mediated reactive pathways, because electron shuttling will simultaneously stimulate Fe(III) reduction and RDX reduction. Reduced electron shuttles will directly transfer electrons to
RDX, but reduced electron shuttles also reduce Fe(III) quickly, which generates a high concentration of reactive Fe(II). While the ultimate end result is RDX transformation in both pathways, it is critical to understand the dominant reaction to better develop the remediation strategy. Although both reaction mechanisms transform RDX, the metabolite distribution is different depending on the dominance of microbial or chemical reactions (Kwon and Finneran, 2008a). It is also likely that direct electron shuttle transformation ( electron shuttles RDX) generates different intermediates than the route from electron shuttles Fe(III) (s) Fe(II) RDX transformation .
The metabolites and intermediates (the former generated by cells and the latter by abiotic reactions, collectively referred to as reaction products) are important because several are equally as toxic as RDX, and the potential for mineralization changes as the reaction products change (Kwon and Finneran, 2008a). In addition, these reaction products are monitored as evidence for in situ biotransformation (by practitioners). Understanding the actual pathway will allow practitioners to know whether the “attempted” remediation strategy is working as planned or whether it is alternate reactions that are driving RDX transformation.
65 While electron shuttling may become a useful strategy for increasing RDX transformation it is difficult to monitor because the “net Fe(II)” often seen as a result of electron shuttling for other contaminants may not be seen with RDX, due to the rapid Fe(II) oxidation by RDX and its reaction products (Gregory et al., 2004). This study used a model Fe(III) reducing culture and a ligand (Ferrozine) that inhibits electron transfer from Fe(II) to RDX to determine if RDX transformation is primarily due to accelerated Fe(II) production, or by direct electron transfer from reduced electron shuttles. This question arose from aquifer material incubations in which
RDX was reduced most readily in the presence of electron shuttles and was linked to
Fe(III) reduction, but the actual mechanism was obscured because “net Fe(II)” did not increase until RDX was near detection limits.
The purpose of this study was to determine the relative extent of direct electron transfer from hydroquinone electron shuttles to RDX versus the proportion from accelerated Fe(III) reduction and Fe(II) electron transfer. We followed contaminated sediment incubations with pure culture and strictly chemical phase experiments to quantify the relative extent of RDX degradation mediated by each process. The data were used to develop a conceptual model where electron shuttling dependent processes were more efficient for RDX transformation than processes that lacked electron shuttles.
MATERIALS AND METHODS
Sediments and Groundwater . Aquifer sediment samples were obtained from the
Picatinny Arsenal in New Jersey. The sediment was collected from below the water table and immediately dispensed into amber glass jars that were then sealed without a headspace. Samples were placed in coolers and shipped via overnight carrier to the laboratory. Anaerobic sediment was homogenized in a N 2 filled glove bag prior to 66 processing for individual experiments. Groundwater from the monitoring well at
Picatinny Arsenal in New Jersey was sampled with a peristaltic pump. Groundwater samples were collected in 1L glass bottles that were then sealed without headspace.
Groundwater samples were stored at 4ºC until processing. The groundwater contained nitrate (0.027 mM), nitrite (0.002 mM), chloride (0.236 mM), bromide (0.002 mM), and sulfate (0.144 mM). The porewater was initially contaminated by RDX (< 1 M).
The aquifer sediment contained 3.9 mmol/kg of bioavailable (hydroxylamine reducible Fe(III) under acidic conditions) Fe(III) (s) .
Sediment Incubations . The aquifer sediments were tested with initial aeration (to oxidize native Fe(II) to Fe(III)) (initial Fe(II) (aq) and Fe(III) (s) concentrations were approximately 90 M and 3.9 mmol/kg, respectively). For initial aeration, the aquifer sediment (50 g each) was aerated with ambient air for 30 minutes and capped with a rubber stopper and an aluminum cap. Then, each serum bottle was flushed with 95:5
(vol/vol) N 2:CO 2 that had been passed over hot copper filings to remove traces of oxygen. Approximately 20 ml of groundwater was dispensed into these bottles using sterile needles and syringes that had been flushed with anoxic gas. The initial concentration of RDX was approximately 1 M. To better detect and quantify RDX loss and reaction product accumulation we added RDX for a final concentration of 60
M. Each bottle had an overlying aqueous layer without forming a slurry. Acetate was added as a sole electron donor at a final concentration of 2 mM. All amendments were made from sterile, anoxic stock solutions and transferred using sterile needles and syringes that had been flushed with anoxic gas. All bottles were incubated in the dark at 18ºC without agitation.
Electron acceptors/shuttles used with the sediments included Sigma Aldrich humic acid (hereafter humic substances (HS)) and AQDS. AQDS and HS were 67 prepared in anoxic 30 mM bicarbonate buffer and dispensed into the 160 mL anaerobic serum bottles for final concentrations of 100 M and 0.15 g/L, respectively.
HS had no direct electron donating capacity (via embedded, reduced Fe(II)). RDX did not adsorb to HS or AQDS.
In order to generate abiotic controls, sediments were autoclaved for 1 h per day for 3 consecutive days (Finneran and Lovley, 2001). All liquid samples were taken with a sterile syringe and needle that had been flushed with anaerobic gas. In order to quantify RDX, its nitroso metabolites, and Fe(II), samples (0.3 ml) were collected periodically via anaerobic syringe and needle. No more than ten samples were taken from any incubation; therefore, final volume of each incubation was approximately 17 mL at the end of the experiments (more than 85% volume remaining).
Medium and culturing conditions. The basal anaerobic medium consisted of (g l 1 unless specified otherwise): NaHCO 3 (2.5), NH 4Cl (0.25), NaH 2PO 4 H 2O (0.6), KCl
(0.1), modified Wolfe’s vitamin and mineral mixtures (each 10 ml l 1) and 1 mL of 1 mM Na 2SeO 4. Soluble Fe(III) citrate (45 mM) was the electron acceptor with the anoxic medium. The medium was buffered using a 30mM bicarbonate buffer equilibrated with 80:20 (vol/vol) N 2:CO 2, as previously described (Finneran et al.,
2003). Culture tubes/bottles were sterilized by autoclaving at 121 °C under pressure for 20 minutes. Acetate (20 mM) was added as the sole electron donor. Standard anaerobic and aseptic culturing techniques were used throughout (Lovley and Phillips,
1988). The medium was sparged with anoxic gases passed over a heated, reduced copper column to remove trace oxygen from the gas line. Culture tube/bottle headspaces were flushed with the same gas mixture and sealed under an anoxic headspace using a thick butyl rubber stopper fastened with an aluminum crimp. 68 Cultures were maintained in 160 mL anaerobic serum bottles; individual experiments utilized different culture tubes or bottles, as described below.
Resting cell suspensions. Resting cell suspensions of Geobacter metallireducens (in the presence and absence of electron shuttles) with RDX were used to determine if the
RDX transformation in aquifer incubations was due to direct electron transfer from anthrahydroquinone disulfonate (AH 2QDS) or reduced HS, or whether Fe(II) was a necessary intermediate electron carrier. Ferrozine, a chemical used in ferrous iron analyses, is a strong ligand that negates Fe(II) electron transfer to the cyclic nitramines.
G. metallireducens was grown anaerobically in freshwater medium with acetate as the sole electron donor and Fe(III) (aq) citrate as the sole terminal electron acceptor. One liter of each individual cell culture was harvested during logarithmic growth phase and centrifuged at 5250 g for 15 minutes to form a dense cell pellet.
Each resultant cell pellet was resuspended in 30 mM bicarbonate buffer under a stream of anoxic gas. The washed cells were centrifuged again at 5250 g for 15 minutes and the resultant biomass was resuspended in 4.0 mL of bicarbonate buffer.
Cells were used within 30 minutes of processing.
Experimental tubes were prepared by sparging approximately 5.0 ml of 30 mM bicarbonate buffer with anoxic N 2:CO 2 (99.9:0.1; pH = 8.2) and sealing the buffer under an anoxic headspace. Electron acceptors were amended from concentrated, anoxic, sterile stock solutions. Electron acceptors incubated with the cells included HS (0.25 g/L), poorly crystalline Fe(III) hydroxide (10 mmol/L),
AQDS (100 M), HS plus poorly crystalline Fe(III) hydroxide, and AQDS plus poorly crystalline Fe(III) hydroxide. Ferrozine reagent (10 mM) was added to one set of cell suspensions as an Fe(II) chelator. RDX and HMX were amended at a final 69 concentration of 46 M and 6 M, respectively, in the resting cell suspension of G. metallireducens . An aliquot (0.3 ml) of the resting cells was added to the sealed pressure tubes to initiate each experiment. Individual tubes were incubated at 30°C.
Samples (0.5 ml) were collected periodically via anoxic syringe and needle.
Stop Flow Kinetic Experiment . In order to measure rapid kinetics of Fe(II) production by AH 2QDS, a stopped flow technique with double mixing option was used. One 10 ml syringe was filled with mixed solutions of Fe(III) (aq) citrate (1 mM)
2+ and Ferrozine (5 mM). Approximately 0.45 mM of Fe preexisted in the Fe(III) (aq) citrate solution bubbled with N 2:CO 2, which was the baseline of the experiment. The other syringe was filled with AH 2QDS (500 M). All stocks were pH 8.2, anoxic and sterile. Each solution in the syringes was injected through a high efficiency mixer.
The reduction kinetics was measured every 2 seconds by a spectrophotometer at an absorbance of 562 nm.
RDX reduction by soluble Fe(II). Biologically reduced iron, soluble Fe(II), was prepared by incubating Geobacter metallireducens in ferric citrate medium (45 mM) .
The soluble Fe(II) was filtered through a 0.2 m sterilized PTFE filter into a pre sterilized, anaerobic serum bottle to remove cells. The concentrations of Fe(II) tested were 1 or 1.5 mM with RDX (35 M). The concentrations of poorly crystalline
Fe(III) oxides (referred to as “Fe Gel”) tested was 5 mmol/L. Ferrozine (5 mM) was amended as the Fe(II) chelator. Incubations were performed at 30°C. Samples were collected periodically via anoxic syringe and needle; samples were filtered through sterile, 0.2 m PTFE filters prior to analyses.
70 RESULTS AND DISCUSSION
RDX Transformation in Contaminated Aquifer Material correlated to Fe(III)
Reduction
RDX was reduced in aquifer material amended with 2 mM acetate; only 20% of the RDX was removed without acetate and sterile controls did not transform any
RDX within 65 days of incubation (Figure 5.1). The results suggest that the availability of an electron donor is a rate limiting factor in degrading RDX in this sediment and indigenous microorganisms utilizing acetate as an electron donor are responsible for RDX biodegradation. AQDS and HS increased the rate of RDX transformation, and the production/loss of the nitroso metabolites (Figures 5.1 and
5.2). Mass balances between RDX reduction and nitroso metabolites produced at day
65 were 0 to 5% in the presence of electron shuttles, suggesting further transformation to other ring cleavage products (e.g., HCHO, MEDINA, NDAB, ammonium, nitrous oxide, methanol, CO 2, etc). AQDS in particular promoted relatively fast RDX removal compared to the alternate treatments (Figure 5.1). MNX barely accumulated in AQDS + acetate amended incubations; DNX and TNX were negligible in AQDS + acetate amended incubations but were present in HS + acetate amended incubations and with acetate alone (Figure 5.2).
Nitroso metabolite production and degradation phenomena were consistent with past pure culture incubations or chemical suspensions with AH 2QDS alone
(Kwon and Finneran, 2008a). RDX was degraded most rapidly in the AQDS + acetate amended incubations, demonstrating that the model quinone accelerates direct
RDX reduction, or Fe(III) reduction mediated RDX transformation. HS also decreased the total timeframe for RDX reduction relative to acetate alone; however, the difference was smaller than that of AQDS + acetate amended incubations. This is likely due to the differences in the redox potential for AQDS and HS and/or the 71 structural differences between AQDS and HS. AQDS is a low molecular mass mono quinone with an easily accessible functional group. Humic substances are undefined molecules that are assemblages of lower molecular mass core functional groups
(Sutton and Sposito, 2005). Humic substances may or may not contain quinone functional groups (Stevenson, 1982), though the proportion of quinone functional groups in purified humic acids has been reported (Scott et al., 1998). There are numerous data supporting faster reaction kinetics between cells and AQDS relative to
HS (Lovley et al., 1996b; Lovley et al., 1998; Lovley et al., 2000; Kwon and
Finneran, 2006, 2008b). Concentrations of humic substances in groundwater and surface water range from 0.1 mg C/L to several hundred mg C/L (Aiken et al., 1985).
We have used 50 mg C/L of HS (approximately 50 weight % C present in HS). We have investigated alternative shuttling compounds for in situ application and the preliminary data indicate that raw (un purified) humic material extracted from mulch
(0.04 0.25 g/L) can stimulate RDX reduction.
Net aqueous Fe(II) was monitored to determine if Fe(III) (s) was being reduced concurrently with RDX reduction. Fe(II) (aq) in solution followed the expected pattern based on electron shuttle + acetate amendment versus acetate alone. All had increases in Fe(II) (aq) , but Fe(II) (aq) increased to the highest concentration in the shortest time with AQDS followed by HS, while slightly less Fe(II) was generated, more slowly, in the acetate alone incubations (Figure 5.3A). When we focused on the first 20 days of the AQDS + acetate amended incubations a unique Fe(II) pattern emerged. The net Fe(II) increased greatly when RDX was at or below 5 M (Figure
5.3B). This was also evident with HS and acetate alone but on a longer time scale
(Figure 5.3A). The data can be interpreted two different ways. The first possible interpretation is that the hydroquinone transferred electrons directly to RDX so quickly that very few hydroquinone electrons were available to reduce Fe(III) (s) to 72 Fe(II) (aq) until RDX reached a threshold level (pathway III in Figure 5.4). The alternate possibility is that the hydroquinone transferred electrons to Fe(III)(s) rapidly, but all Fe(II) generated reduced RDX via previously described mechanisms (e.g., surface bound Fe(II)) (Gregory et al., 2004; Kwon and Finneran, 2006; Larese
Casanova and Scherer, 2008) (pathway IV in Figure 5.4). In the latter scenario the
“net Fe(II)” in solution would be minimal until RDX was below a threshold concentration.
Figure 5.4 is a conceptual model of the available pathways for a typical
Fe(III)/electron shuttle reducing cell for reducing RDX. Pathways I and II are negligible when electron shuttles are present; electron shuttle reduction is rapid and pathways III and IV will predominate. However, it is unclear whether pathway III or
IV is more relevant to RDX degradation, as both have been shown previously (Kwon and Finneran, 2006, 2008a, 2008b). Pure culture and strictly chemical phase experiments were designed and performed to address this question, as described below.
Kinetics of AH 2QDS Electron Transfer to Fe(III)
Electron transfer from AH 2QDS to any form of Fe(III) has been referred to as
“instantaneous” in the Fe(III) reducing literature (Benz et al., 1998; Kwon and
Finneran, 2006). However, even “instantaneous” has a rate, albeit a very fast one. A stop flow kinetics spectrophotometer was used to quantify the actual rate of electron transfer from AH 2QDS to Fe(III) (aq) citrate. A soluble form of Fe(III) was necessary for this experimental system; however, the data can be used as a general estimate of rates of electron transfer from AH 2QDS to Fe(III). Even if our kinetics based on the soluble form of Fe(III) rather than solid phase Fe(III) are overestimates, Fe(III) (s) reduction by hydroquinones is likely still much faster than RDX reduction. 73 Our previous data determined that the maximal rate of electron transfer from
1 AH 2QDS to RDX was approximately 0.17 mol RDX hour , at pH 8.2. The pH is critical because hydroquinones in general are better reductants at alkaline pH
(Uchimiya and Stone, 2006; Kwon and Finneran, 2008a). We tested identical pH conditions with AH 2QDS and Fe(III) (aq) citrate in a stop flow device (fast mixing system, no diffusion controlled) that can measure rates of Fe(III) (aq) reduction (by
Fe(II) Ferrozine color formation) on a scale of milliseconds. 1.0 mM Fe(III) (aq) citrate was completely reduced by AH 2QDS in 100 seconds, with an apparent first order rate
1 of 11 mol Fe(III) (aq) reduced second (Figure 5.5). Therefore the rate of electron
5 transfer from AH 2QDS to Fe(III) (aq) is approximately 10 times faster than the rate of
AH 2QDS electron transfer to RDX. This clearly argues for pathway IV (Figure 5.4) as the dominant pathway – given that every electron from AH 2QDS should be oxidized by Fe(III) (s) rather than RDX. Even if the in situ rates of AH 2QDS
Fe(III) (s) electron transfer are much slower than this model system it is still significantly faster than AH 2QDS RDX electron transfer.
While the kinetics argument is straightforward between the two pathways, our past data (Kwon and Finneran, 2006) suggest that some fraction of electrons from
AH 2QDS will directly reduce RDX even when Fe(III) (s) is present, irrespective of pH.
Kinetics alone cannot always determine the actual pathway when there is competition between two mechanisms, especially when the overall reaction is a mixed biological abiotic pathway. We therefore performed additional experiments with a compound that specifically blocks Fe(II) mediated RDX reduction to determine if electrons from
AH 2QDS or reduced humic substances directly reduce RDX even when Fe(III) (s) is present, described below.
74 Delineating Electron Transfer Pathways using Ferrozine
The ferrous iron ligand Ferrozine is a hexadentate ligand that strongly coordinates around Fe(II), forming a soluble, stable and unreactive complex. It is most often used in analytical chemistry of ferrous iron (Stookey, 1970). However, when present in slight excess of total Fe(II) it can limit or completely inhibit Fe(II) mediated reduction. The concentration of Ferrozine used in these experiments did not interfere with cellular metabolism (data not shown); therefore all decreases in RDX reduction in the presence of Ferrozine were due to blocking electron transfer from
Fe(II) RDX. Abiotic suspensions of Fe(II) Ferrozine complexes did not react with
RDX over a period of days, but Fe(II) alone (in the presence of reactive Fe(III) surfaces) reacted very quickly (data not shown). The role of Ferrozine to increase dissimilatory iron reduction by alleviating Fe(II) inhibition has been also suggested
(Royer et al., 2002a), as Fe(II) adsorption onto Fe(III) (hydr)oxide and microbial cell surfaces inhibits further iron reducing activity (Roden and Zachara, 1996; Urrutia et al., 1998). However, Fe(II) complexation by Ferrozine enhanced hematite reduction in the case of excess Fe(II) accumulation in the system with incubation times that were much longer than those presented here (Royer et al., 2002b).
Resting cell suspensions of G. metallireducens were selected to represent
Fe(III) (s) /AQDS/RDX reduction because the culture is known to reduce all compounds involved, including RDX and its metabolites to a small extent (Kwon and
Finneran, 2006). Three different suspensions were run: the first with cells + Fe(III) (s) in the presence and absence of Ferrozine – no electron shuttles (Figure 6A), the second with cells + AQDS + Fe(III) (s) in the presence and absence of Ferrozine
(Figure 5.6B), and the last with cells + HS + Fe(III) (s) in the presence and absence of
Ferrozine (Figure 5.6C).
75 RDX reduction was limited when electron shuttles were absent, which is consistent with previous data (Kwon and Finneran, 2006, 2008b) (Figure 5.6A).
However, even with the limited RDX reduction there was a quantifiable difference between the Ferrozine amended suspensions and the suspensions without Ferrozine.
Aqueous Fe(II) accumulated in the presence of Ferrozine, which was expected because the pathway for Fe(II) oxidation was cut off.
The data with electron shuttles present are much more clear and demonstrate that RDX is concurrently reduced by electron shuttles directly and by Fe(II) generated during electron shuttle mediated Fe(III) reduction (Figure 5.6B and 5.6C). Both HS and AQDS (+ acetate) amended suspensions fully reduced RDX with minimal accumulation of aqueous Fe(II) when Ferrozine was absent. However, Ferrozine limited RDX reduction in both suspensions but did not completely inhibit it (Figures
5.6B and 5.6C). Ferrozine functionally excludes Fe(II) oxidation by RDX, but not
Fe(III) reduction or electron shuttle reduction. Figure 5.6A and past data (Kwon and
Finneran, 2008b) demonstrate that RDX reduction directly by cells is negligible, so any RDX reduction present when Fe(II) electron transfer was blocked was due to direct electron transfer from either AH 2QDS or reduced HS. Relative amounts (%) of
RDX loss due to the pathway III and IV were compared. Assuming pathways I and II were negligible, pathway III accounted for approximately 40% of the RDX reduced with either AQDS or HS, respectively. The remainder of the RDX reduction was mediated by pathway IV. Ferrous iron accumulated in both AQDS and HS (+ acetate) amended suspensions when Fe(II) oxidation was blocked, as expected. Reaction products other than the nitroso intermediates were quantified over time including
MEDINA, NDAB, formaldehyde, nitrous oxide, nitrite, and ammonium. The reaction products were consistent with what has been previously identified with total iron and electron shuttles, though adding Ferrozine shifted the overall mass balance to more 76 residual RDX and nitroso metabolites. Carbon mass balances ranged from 83 95% and nitrogen mass balances ranged from 82 97%, respectively.
Electron shuttling is important from both a fundamental and applied perspective for a number of reasons. First, the fundamental data establish that there are multiple, simultaneous pathways for RDX transformation in the presence of electron shuttles. Numerous pathways have been proposed for RDX biotransformation under anoxic and oxic conditions (Hawari et al., 2000a; Crocker et al., 2006), and these data support that no single pathway is clear cut in the presence of iron and extracellular electron shuttles. From the applied perspective this shows that electron shuttling can accelerate RDX degradation in situ (Figure 5.1) and that degradation will ensue via electron shuttles in the presence and absence of Fe(III) (s) .
It also suggests that soluble Fe(II) in groundwater may not be an appropriate indicator of Fe(III) reduction in the presence of contaminants that will oxidize Fe(II). The “net
Fe(II)” will not increase even though Fe(III) (s) is being reduced unless RDX (or another oxidizing contaminant) falls below a threshold concentration. One question that seems to arise as a result of this and other work is that given so many pathways for RDX degradation – why is contamination so widespread? One possibility is limited mass transfer of RDX from the adsorbed phase to the dissolved phase, which will limit all of these reactions. Another possibility is lack of anoxic conditions in the areas of highest contamination. This and other work demonstrates that anoxic conditions favor rapid RDX transformation. If the dominant contamination is within oxic zones, many of these reactions will not be favored.
To reinforce the strategy of electron shuttling we also tested the quadruple nitro group cyclic nitramine explosive HMX in identical cell suspensions under identical conditions. Past data (Kwon and Finneran, 2008b) have demonstrated that both AH 2QDS and reduced HS directly reduce HMX as well as reduced shuttle 77 Fe(III) (s) Fe(II) HMX. The rates are slower because of the additional nitro group, but all reactions with RDX and electron shuttles have been mirrored by reactions with HMX (data not shown). HMX was slightly reduced by cells +
Fe(III) (s) , but Ferrozine had the same effect as it did with RDX. AQDS and HS increased the rate and extent of HMX degradation. As with RDX, Ferrozine limited the amount of HMX reduced but did not completely inhibit HMX reduction. Fe(II) accumulated in all Ferrozine amended suspensions, demonstrating that both electron transfer pathways were acting simultaneously. We monitored the nitroso metabolites of HMX and all were produced (data not shown). Metabolites downstream of the nitroso compounds were not monitored for HMX.
Fe(II) (aq) -Mediated RDX Reduction without Cells Present
RDX reduction by the filtered, soluble Fe(II) in the presence or absence of the poorly crystalline Fe(III) hydroxide and/or Ferrozine was investigated since surface complexed Fe(II) or structural Fe(II) may lead to reduction of RDX by an alternative route . The catalytic activity of reactive Fe(II) precipitates (e.g., green rust, mackinawite, and magnetite), as well as Fe(II) adsorbed on (hydr)oxides, have been shown to reduce a number of environmentally relevant compounds (Klausen et al.,
1995; Charlet et al., 1998; Buerge and Hug, 1999; Liger et al., 1999; Amonette et al.,
2000; Charlet et al., 2002; Vikesland and Valentine, 2002; Klupinski and Chin, 2003;
O'Loughlin et al., 2003; Strathmann and Stone, 2003). Unaltered poorly crystalline
Fe(III) hydroxide (5mM) alone or soluble Fe(II) alone in cell free incubations did not reduce RDX within 80 hours (Figure 5.7). Poorly crystalline Fe(III) hydroxide with addition of soluble Fe(II) slowly reduced RDX (Figure 5.7) possibly by providing a surface for catalyzing the electron transfer. The catalytic activity of surface bound
Fe(II) on the poorly crystalline Fe(III) hydroxide may be the result of complexation of 78 Fe(II) with surface hydroxyl groups (Buerge and Hug, 1998; Strathmann and Stone,
2002a, 2002b) and adsorbed Fe(II) at the surface may form inner sphere bonds, which can increase the electron density of the adsorbed Fe(II) (Stumm, 1992; Charlet et al.,
1998). Incubations with 1.5 mM soluble Fe(II) reduced RDX faster than with 1 mM soluble Fe(II) (Figure 5.7). Adding Ferrozine to soluble Fe(II) plus Fe(III) hydroxide incubations inhibited RDX reduction (Figure 5.7) suggesting that the Ferrozine binds
Fe(II) faster than the Fe(II) will bind Fe(III) hydroxide surfaces to promote surface catalyzed reduction. Although several types of Fe(II) phases (e.g. magnetite) capable of reducing RDX can be present to some extent in these abiotic incubations, adsorbed
Fe(II) (aq) is expected to be the most dominant RDX reducing species.
The initial Fe(III) phase and the resulting Fe(II) phase(s) will have a significant effect on the rate and extent of RDX transformation. The iron phases will also impact the extent to which electron shuttling influences RDX transformation, by changing the proportion of electrons that are transferred to Fe(III), versus those directly transferred to RDX. While the experimental systems reported here do not account for all possible Fe(III) and Fe(II) phases, they do discriminate between direct electron transfer from the shuttles (to RDX), or RDX transformation mediated by electron shuttles + iron. This is a first step to understanding the complex biogeochemical interactions that degrade RDX when electron shuttles are naturally present, or added in engineering applications.
CONCLUSIONS
The results of this investigation provide useful information for developing strategies for RDX biodegradation based on electron shuttling. While this is still a maturing technology these data provide evidence that RDX reduction is in fact accelerated in batches mimicking in situ conditions by adding AQDS or HS. The 79 extent of nitroso metabolite accumulation is lower with electron shuttles, and the
Fe(II) generated by electron shuttles and Fe(III) reduction contributes to RDX degradation.
The number of simultaneous reactions that are stimulated by electron shuttles increases the extent of RDX transformation. First, the microorganisms stimulated and enriched may directly reduce RDX. Several past reports have identified organisms that are known Fe(III) reducers as direct RDX reducers (Kwon and Finneran, 2006,
2008b). We used Geobacter within these experiments for ease of manipulation, but alternate microbial communities that reduce Fe(III) and electron shuttles will most likely be stimulated by this strategy. If these organisms also directly reduce RDX quickly, then the biomass enriched will benefit ongoing natural attenuation once the engineered strategy ceases. Second, we have shown that two abiotic processes are attacking RDX at the same time. These are conceptualized in Figure 4 as pathways
III and IV. This is beneficial because it means electron shuttle mediated strategies do not need to rely solely on pathway III or IV. As the iron mediated pathway diminishes the direct electron shuttling pathway may become dominant, assuming appropriate environmental conditions.
The kinetic data with AH 2QDS and Fe(III) are important to RDX degradation but also to the broader scientific community studying electron shuttling. The previously reported “instantaneous” Fe(III) reduction now has a rate constant that can be used for predictive purposes in future Fe(III) reduction studies. This will also help all bioremediation applications in which electron shuttling is engineered into the system or in which it is part of the natural attenuation capacity of the environment.
We can use these kinetics data as a starting point to determine the rates of Fe(III) reduction versus other processes catalyzed by electron shuttles.
80
Figure 5.1. RDX reduction in contaminated aquifer material from the Picatinny Arsenal that has been amended with the electron shuttles AQDS (100 M) or purified humic substances (HS) (0.15 g/L), or which has only been amended with the electron donor acetate (2 mM). The aquifer material was aerated before beginning the experiments. Results are the means of triplicate analyses and bars indicate one standard deviation
81
Figure 5.2. The production and reduction of MNX (2A), DNX (2B) and TNX (2C) in contaminated aquifer material from the Picatinny Arsenal that has been amended with the electron shuttles AQDS (100 M) or purified HS (0.15 g/L), or which has only been amended with the electron donor acetate (2 mM). The aquifer material was aerated before beginning the experiments. Results are the means of triplicate analyses and bars indicate one standard deviation 82
Figure 5.3. Fe(II) production (3A) in contaminated aquifer material; the comparison of RDX reduction and Fe(II) production (3B) for the first 20 days in contaminated aquifer material. Results are the means of triplicate analyses and bars indicate one standard deviation
83
Figure 5.4. Conceptual model of multiple electron transfer mechanisms in the presence of both Fe(III) and extracellular electron shuttles (EES); Each number (I IV) indicates different electron transfer pathways (per the discussion section)
84
Figure 5.5. Rapid production of Fe(II) by AH 2QDS. Fe(II) concentrations was measured every 2 seconds by Stopped flow technique. The solid line indicates measured values and the dashed line indicates the calculated rate of Fe(II) (aq) production
85
50 0.10
40 0.08 cells+Ac+FeGel cells+Ac+FeGel+FZ 30 0.06 M) no cells+FeGel+FZ 6A cells+Ac+FeGel (Fe(II)) 20 0.04 RDX (
Fe(II)(mM) cells+Ac+FeGel+FZ (Fe(II)) no cells+FeGel+FZ (Fe(II)) 10 0.02
0 0.00 0 10 20 30 40 50 60 50 0.30 time (h)
0.25 40 cells+Ac+AQDS+FeGel 0.20 30 cells+Ac+AQDS+FeGel+FZ M) no cells+AQDS+FeGel+FZ 0.15 6B cells+Ac+AQDS+FeGel (Fe(II)) 20 RDX (
Fe(II)(mM) cells+Ac+FeGel+AQDS+FZ (Fe(II)) 0.10 no cells+AQDS+FeGel+FZ (Fe(II)) 10 0.05
0 0.00 0 10 20 30 40 50 60 50 1.0
40 0.8 cells+Ac+HS+FeGel cells+Ac+HS+FeGel+FZ 30 0.6 M) no cells+HS+FeGel+FZ 6C cells+Ac+HS+FeGel (Fe(II)) 20 0.4 Fe(II)(mM) RDX ( cells+Ac+HS+FeGel+FZ (Fe(II)) no cells+HS+FeGel+FZ (Fe(II)) 10 0.2
0 0.0 0 10 20 30 40 50 60 time (h)
Figure 5.6. RDX reduction and Fe(II) accumulation with or without Ferrozine (FZ) reagents in the presence of poorly crystalline Fe(III) hydroxide (FeGel) only (6A), FeGel plus AQDS (6B), or FeGel plus humic substances (HS) (6C) in cell suspension incubations of G. metallireducens . Arrows indicate differences of the extent of RDX reduction in the presence or absence of Ferrozine. The closed symbols correspond to RDX, while the open symbols to Fe(II). Results are the means of triplicate analyses and bars indicate one standard deviation
86
Figure 5.7. RDX reduction by aqueous Fe(II) added to suspensions of poorly crystalline Fe(III) oxide (FeGel), in the presence or absence of the ligand Ferrozine (FZ) to block Fe(II) electron transfer. Fe(II) was added at 1 mM or 1.5 mM, and Fe(III) was always present in excess (~5 mmol/L). Ferrozine was added at concentrations stoichiometric to the added Fe(II). Results are the means of triplicate analyses and bars indicate one standard deviation
87 CHAPTER 6
HEXAHYDRO 1,3,5 TRINITRO 1,3,5 TRIAZINE (RDX) MINERALIZATION
AND BIOLOGICAL PRODUCTION OF 4 NITRO 2,4 DIAZABUTANAL
(NDAB) IN THE PRESENCE OF EXTRACELLULAR ELECTRON
SHUTTLING COMPOUNDS AND A UNIQUE FE(III) REDUCING
MICROBIAL COMMUNITY 6
INTRODUCTION
Hexahydro 1,3,5 trinitro 1,3,5 triazine (RDX) biodegradation is a promising strategy for decontamination of Department of Defense (DoD) or private facilities that use or produce live munitions (Adrian et al., 2003; Adrian and Arnett, 2004;
Sherburne et al., 2005; Thompson et al., 2005; Meyers et al., 2007). Nitro groups on the RDX molecule can be microbially or chemically reduced, destabilizing the ring structure and leading to spontaneous ring cleavage (McCormick et al., 1981; Hawari,
2000). Different transformation products (e.g., nitroso compounds, methylenedinitramine (MEDINA), 4 nitro 2,4 diazabutanal (NDAB), nitrous oxide
+ (N 2O), nitrite (NO 2 ), ammonium (NH 4 ), formaldehyde (HCHO), formic acid
(HCOOH), and methanol (CH 3OH)) have been quantified under varying conditions.
Some of these products are then available for further metabolism, as electron donors or nitrogen sources, if the appropriate microbial community is present to ultimately mineralize the carbon and/or nitrogen metabolites (Hawari et al., 2000a). One major limitation of RDX bioremediation is that it often leads to accumulated intermediates, without the compounds ultimately being mineralized to CO 2 or CH 4. The limited mineralization may be a result of targeting microbial communities that cannot further 6 Reproduced in part with permission from Kwon, M.J. and K.T. Finneran. Environmental Science and Technology , “Submitted for Publication”
88 oxidize the intermediates as carbon substrates either to gain energy or via co metabolic reactions.
Several enrichment cultures and/or pure cultures have been used to investigate
RDX biodegradation. Data have been reported for aerobic conditions (Seth Smith et al., 2002; Fournier et al., 2005; Thompson et al., 2005) , nitrate reducing conditions
(Freedman and Sutherland, 1998), sulfate reducing conditions (Boopathy et al., 1998), methanogenesis (Adrian et al., 2003), and acetogenesis (Adrian and Arnett, 2004;
Sherburne et al., 2005) . Bradley et al reported native in situ mineralization in one contaminated sediment associated with Mn(IV) reducing conditions (Bradley and
Dinicola, 2005); Mn(IV) is reduced by organisms that also reduce Fe(III). The work presented focuses here on dissimilatory Fe(III) reduction; the data demonstrate that this is a primary process promoting in situ RDX biodegradation. In addition, stimulating Fe(III) reduction via extracellular electron shuttling compounds increased the overall extent of mineralization, which suggests the process can be accelerated.
Fe(III) reduction is an important biogeochemical process due to the abundance of ferric minerals (e.g., ferrihydrite, goethite, or phyllosilicate clay minerals) in subsurface environments and the number of reactions carried out by the associated microorganisms (Lovley, 1995; Roden et al., 2000; Zachara et al., 2001; Kostka et al.,
2002). Fe(III) reducing microorganisms are ubiquitous in subsurface systems; therefore, it is a metabolism that can be stimulated in all environments (Coates et al.,
1998; Lovley, 2000b). Fe(III) reducing microorganisms also reduce extracellular electron shuttling compounds such as humic substances (HS) or model extracellular quinones. Electron shuttling has been investigated for explosives transformation by fermentative microorganisms (Borch et al., 2005; Bhushan et al., 2006) or respiratory microorganisms (Kwon and Finneran, 2006, 2008a, 2008b). Electron shuttling compounds can promote contaminant reduction by accepting electrons in microbial 89 respiration and abiotically transferring the electrons to contaminants (Lovley, 2000b).
The electron shuttles are re oxidized and available again for microbial respiration; this cycling allows a relatively low concentration to continuously promote these reactions.
While environmentally relevant electron shuttles need to be identified (i.e. AQDS is not reasonable for field applications), all data suggest it is a good strategy for accelerating iron dependent reactions.
Recent data in our lab and by others suggest that mixed biotic abiotic reactions involving Fe(III) reducing microorganisms, electron shuttling compounds, and reactive Fe(II) degrade RDX effectively (Gregory et al., 2004; Kwon and
Finneran, 2006; Kim and Strathmann, 2007; Larese Casanova and Scherer, 2008;
Naja et al., 2008). However, what remains unknown is: a) the extent to which electron shuttling will increase mineralization in RDX contaminated aquifer material, b) what transformation products arise specifically due to this metabolism as opposed to alternate microbial chemical pathways, and c) what microorganisms will catalyze these reactions in the presence and absence of added electron shuttles.
The reaction product NDAB has been a focus in several previous studies because it was identified as a “biological metabolite” only under aerobic conditions and an “abiotic intermediate” under anoxic conditions (Fournier et al., 2002; Hawari et al., 2002; Balakrishnan et al., 2003; Bhushan et al., 2003b; Fournier et al., 2004a;
Just and Schnoor, 2004; Fournier et al., 2005). This made pathway determinations seemingly straightforward because NDAB was produced in a distinct manner given the prevalent geochemical conditions. However, our data demonstrate that NDAB does in fact accumulate under anoxic conditions, due to Fe(III) reducing microbial activity. Sediment experiments and pure cultures were used to confirm these data, and NDAB is a distinct biological metabolite, while formaldehyde is the dominant abiotic intermediate. To the best of our knowledge this is the first report of biological 90 NDAB formation under anoxic conditions.
The purpose of this study was to quantify the extent to which electron shuttles could stimulate RDX mineralization, to link RDX transformation to Fe(III) reduction, and to characterize the Fe(III) reducing microbial community that developed as it was transformed and ultimately mineralized. Contaminated subsurface material was used to investigate the product distribution of RDX degradation via extracellular electron shuttling compounds and to determine the dominant biogeochemical process (so called “terminal electron accepting process” or TEAP) that promotes RDX biodegradation. The data demonstrated that electron shuttles added to contaminated aquifer material will stimulate RDX mineralization (using U [14 C] RDX) relative to electron donor amendment alone (the typical strategy), and that NDAB can be produced biologically under anoxic conditions, both novel results. The microbial community that developed was unique, in that the “typical” (i.e. Geobacteraceae )
Fe(III) reducing microorganisms were not dominant.
MATERIALS AND METHODS
Sediment and Groundwater . Aquifer sediment samples were obtained from the
Picatinny Arsenal in New Jersey. The sediment was collected from below the water table and immediately dispensed into amber glass jars that were then sealed without a headspace. Samples were placed in coolers and shipped via overnight carrier to the laboratory. Anaerobic sediment was homogenized in a N 2 filled glove bag prior to processing for individual experiments. Groundwater from the monitoring well at
Picatinny Arsenal was sampled with a peristaltic pump. Groundwater samples were collected in 1L glass bottles that were then sealed without headspace. Groundwater samples were stored at 4ºC until processing. The groundwater contained nitrate
(27 M), nitrite (2 M), chloride (236 M), bromide (2 M), and sulfate (144 M). The 91 porewater was initially contaminated by RDX (~ 1 M). The aquifer sediment contained 4.7 mmol/kg of bioavailable (hydroxylamine reducible Fe(III) under acidic conditions) Fe(III).
Sediment Incubations. The aquifer sediments were tested with initial aeration to oxidize native Fe(II) to Fe(III). For initial aeration, the aquifer sediments were aerated with ambient air for 30 minutes. Approximately 50g of sediment and 30ml of groundwater were dispensed into 70ml serum bottles in an N 2 filled glove bag that were then sealed with a thick butyl rubber stopper. After removal from the glove bag, the headspace of each bottle was flushed with 95:5 (vol/vol) N 2:CO 2 that had been passed over hot copper filings to remove traces of oxygen. Approximately 55 M of
RDX was added to the aquifer sediment incubations because the initial concentration was too low to quantify significant losses of the parent compound or accumulation of metabolites. All amendments were made from sterile, anaerobic stock solutions.
Acetate was added as a sole electron donor at a final concentration of 10 mM.
Although aquifer material already contained nitrate and sulfate, additional nitrate (0.5 mM) and sulfate (1.5 mM) were amended to each bottle to increase the capacity for different terminal electron accepting processes. Electron shuttles used with the sediments included HS (0.15 g/L) and anthraquinone 2,6 disulfonate (AQDS) (100
M). All subsequent amendments or transfers were made using sterile needles and syringes that had been flushed with anaerobic gas.
In order to generate abiotic controls, sediments were autoclaved for 1 h per day for 3 consecutive days (Finneran and Lovley, 2001). All bottles were incubated in the dark at 18ºC without agitation. Each bottle had an overlying aqueous layer without forming a slurry. Samples (0.6 ml) were collected periodically via anoxic syringe and needle. To minimize sampling volume, glass inserts (250 L Glass LVI 92 Flat Bottom; Laboratory Supply Distributors, NJ) were used in the autosampler vials.
Nitrous oxide was monitored by headspace analysis. The aqueous phase pH was measured at each time point in a glove bag. Approximately 1 g of sediment sample was collected at each time point (after headspace samples were analyzed) in an anoxic glove bag to measure total bioavailable iron, which was measured for all incubations at each time point. All liquid and gas samples were taken with a sterile syringe and needle that had been flushed with anaerobic gas, and liquid samples were filtered through 0.2 m PTFE filters (PALL Sciences) prior to analyses. All experiments were performed in triplicate.
Sediment Incubations with “Confidential Sediment #2”. Additional aquifer material was obtained from a military area (the location was confidential at the request of the stakeholder) and similar experiments to the Picatinny sediment series were run. The data are presented as part of Table S1 as evidence for anaerobic, biological NDAB formation. Concentrations of nitrate, sulfate, and bioavailable
Fe(III) in the aquifer material were 17 M, 540 M, and 3.9 mmol/kg, respectively.
Initially 37.5 M of acetate (sub stoichiometric) as an electron donor was amended and 2mM of acetate was re amended at day 34 as RDX was not reduced with the lower concentration of acetate. The pH in was 7.5. Other experimental conditions were same as the method described in the main text.
Resting cell suspension . Resting cell suspensions of Shewanella oneidensis MR 1
(in the presence and absence of electron shuttles) with MNX were used to determine if NDAB production was due to direct electron transfer from anthrahydroquinone disulfonate (AH 2QDS) to MNX, or whether cellular activity was a necessary. Resting cell suspension of S. oneidensis was prepared as described previously (Kwon and 93 Finneran, 2008b). Experimental tubes were prepared by sparging approximately 5.0 ml of 30 mM bicarbonate buffer with anoxic N 2:CO 2 (pH = 7.2 or 8.2) and sealing the buffer under an anoxic headspace. AQDS (100 M) was amended as an electron shuttle with the cells and lactate (20 mM). AH 2QDS (100 M) was amended as an electron donor in cell free incubations. MNX was amended at a final concentration of
40 M in the resting cell suspension of S. oneidensis . An aliquot (0.3 ml) of the resting cells was added to the sealed pressure tubes to initiate each experiment.
Individual tubes were incubated at 30°C.
In order to investigate kinetics of RDX degradation and metabolite production
(in particular NDAB production), the resting cell suspension of MJ1 was prepared as described above. MJ1 was grown anaerobically in freshwater medium with glucose
(10 mM). RDX was amended at the concentration of 65 M in the resting cell suspension of MJ1. Other experimental condition was same as above.
Mineralization with [ 14 C] RDX . Uniformly radiolabeled (U [14 C]) RDX
14 (concentration: final radioactivity of 36,000 dpm/ml) was amended at day 32. CO 2
14 14 and CH 4 were monitored by analysis of headspace samples (1ml). H CO 3 was
14 used to establish CO 2 partitioning between the liquid and gas phase, and was
14 factored in to the final mineralization; the partition coefficient (total dpm H CO 3
14 recovered as CO 2/total dpm added) was 0.013. All experiments were performed in triplicate.
Microbial Community Analysis. Approximately 1 g of sediment was collected from each bottle at every time point using a sterile metal spatula in an anoxic glove bag and dispensed into sterile micro centrifuge tubes. Samples were frozen at –80ºC until the experiments were completed. Total genomic DNA was extracted using a Fast DNA 94 SPIN for Soil Kit (MP Biomedicals, LLC.) with a bead beating apparatus according to manufacturer’s directions. Extracted DNA was amplified using universal
Eubacterial primers, 338 Forward (338F) (Amann et al., 1995) and 907 Reverse
(907R) (Lane et al., 1985). PCR reaction mixtures included primer sets (1 M each), deoxynucleoside triphophates (dNTPs) (200 M), PCR 10X buffers (1X), MgCl 2 (2.5 mM), reaction template (extracted DNA), and Taq polymerase (2.5U). The reaction mixture volume for each template was 100 l. The PCR mixtures were sterilized by
UV radiation for 20 minutes prior to the addition of template and Taq polymerase.
Templates were amplified using a thermocycler (an initial denaturation step at 94°C for 4 min, followed by 35 cycles of 94°C (30 s), 50°C (30 s), and 72°C (45 s), with a final extension at 72°C for 7 min).
The initial PCR product was used to generate clone libraries using a TOPO
TA cloning kit (Invitrogen) according to the manufacturer’s specified instructions. At least 50 clones from the clone library were selected for restriction enzyme analysis.
Cloned DNA was re amplified using clone specific primers (0.4 M each), M13F (5’
GTAAAACGACGGCCAG 3’) and M13R (5’ CAGGAAACAGCTATGAC 3’)
(Invitrogen), dNTPs (350 M), PCR 10X buffers (1X), MgCl 2 (2.5 mM), reaction template (individual clone applied directly into reaction mixture), and Taq polymerase
(2.5U). The reaction mixture volume for each template was 50 l. The thermocycling program included an initial step of cell lysing and nuclease inactivation at 94°C for 10 min, followed by 30 cycles of 94°C (1 min), 55°C (1 min), and 72°C (1.5 min) with a final extension at 72°C for 10 min). DNA (12 L) from M13 PCR was digested with two restriction enzymes, Hha I and Msp I (10U each) (New England Bio), 10X NE
Buffer (1X), and BSA (100 g/mL) for 16 hours at 37ºC and 20 minutes at 60ºC.
Restriction patterns were visualized on a 3% Metaphor agarose gel (Lonza, Rockland,
Maine) using a UV box Gel Doc system, and unique clones were purified using a 95 QIAquick PCR Purification Kit (QIAGEN Inc.). Purified M13 PCR products were sequenced at the University of Illinois automated sequencing facility with the M13 forward primer. These sequences were analyzed versus 16S rDNA in the publicly available database GenBank using BLASTN and SIMILARITY RANK algorithms.
Representative sequences were aligned with the reference sequences using Clustal W
(Thompson et al., 1994). The phylogenetic trees were constructed using the maximum likelihood method with Geneious 3.7.0 (Biomatters Ltd.) software.
Nucleotide sequence accession number. Sequences identified in this study were deposited in the GenBank database under the accession numbers EU826721 –
EU826755.
RESULTS AND DISCUSSION
RDX reduction and metabolite distribution
RDX was not reduced or transformed in heat sterilized incubations, or in the absence of acetate (Figure 6.1A). Native electron donors were limited in this aquifer material and all activity required acetate amendment (10 mM). The primary differences between acetate alone and acetate + shuttles were related to the rate and extent of reaction product formation, the products formed, and eventual mineralization (highest with electron shuttles). RDX was reduced in all acetate amended incubations, and the electron shuttles only slightly increased the extent removed at each time point (Figure 6.1A).
The nitroso compounds accumulated to a lesser extent in the presence of electron shuttles, which is consistent with previous data (Kwon and Finneran, 2006,
2008a, 2008b) (Figure 6.1B D). The nitroso compounds are toxic and accumulation is a concern (Pitot and Dragan, 1996; Meyer et al., 2005). Several reports indicate that 96 they are unstable and will spontaneously degrade (Oh et al., 2001; Sheremata et al.,
2001; Gregory et al., 2004). TNX has been reported to degrade by an uncharacterized autocatalytic mechanism (Larson et al., 2005); however, TNX accumulation has also been reported in pure phase, mixed culture and sediment incubations (Gregory et al.,
2004; Larese Casanova and Scherer, 2008). Nitroso metabolite accumulation indicates that the initial steps mediated by the microorganisms with and without electron shuttling compounds was mainly nitro group reduction rather than denitration. Electron shuttles increased the rate of nitroso group transformation, which was shown previously (Kwon and Finneran, 2006). While autocatalytic decomposition is one pathway for nitroso group degradation these data and other recent reports with sulfate and carbonate green rusts (Larese Casanova and Scherer,
2008), Fe(II) ligand complexes (Kim and Strathmann, 2007), and zero valent iron
(Naja et al., 2008) suggest that abiotic reduction accelerates nitro reduction and eventual ring cleavage.
+ The ring cleavage metabolites HCHO, MEDINA, NDAB, NH 4 , N 2O, and
NO 2 were analyzed (Figure 6.1E and 6.2). HCHO and MEDINA were only produced in AQDS amended incubations; both decreased by day 25 (Figure 6.2A and 6.2B).
NDAB is discussed below. Ammonium was the most dominant inorganic nitrogen intermediate in the presence of electron shuttles. Ammonium began to increase after
20 days in acetate amended incubations and accumulated more in the presence of
electron shuttles (Figure 6.2C). Neither nitrite (NO 2 ) nor nitrous oxide (N 2O) was detected in any of the incubations (data not shown). However, our previous study
(Kwon and Finneran, 2008a) showed that G. metallireducens can transform nitrite, produced from denitration of RDX, to ammonium and/or N 2O. Even with the different Fe(III) reducing microbial community it is likely that nitrite produced from
97 RDX reduction was rapidly transformed to ammonium and/or N 2O by biological processes such as nitrite reductase, with further transformation of N 2O.
Biological NDAB production by Fe(III)-Reducing Microorganisms
NDAB increased to approximately 10 M in electron shuttle amended incubations and remained stable for the rest of the sampling time (Figure 6.1E).
NDAB has been reported as an aerobic ring cleavage intermediates during alkaline hydrolysis of RDX (Balakrishnan et al., 2003), during phytophotolysis of RDX (Just and Schnoor, 2004), during photolysis of RDX in aqueous solution (Hawari et al.,
2002), and during aerobic biodegradation of RDX (Fournier et al., 2002; Bhushan et al., 2003b; Fournier et al., 2004a; Fournier et al., 2005). In fact, a recent paper suggests that NDAB is an “aerobic” RDX intermediate (Jackson et al., 2007). The current incubations were performed under strict anoxic conditions with neutral pH
(~7.0) in the dark. NDAB production was unanticipated but demonstrates that there are several possible degradation routes for RDX.
NDAB production in the current study was likely due to the Fe(III) reducing microbial community present in the sediment (see section below). Bhushan et al.
(Bhushan et al., 2003b) and Fournier et al. (Fournier et al., 2002) reported that
Rhodococcus sp. DN22 (Phylum Actinobacteria ) used RDX as a nitrogen source and consequently produced NDAB. It was hypothesized that two denitration steps via two single electron transfers to RDX, mediated by a cytochrome P 450, leads to spontaneous hydrolytic ring cleavage and further decomposition to produce NDAB
(Bhushan et al., 2003b). Though it is possible that some Actinobacteria present in the sediment were responsible for NDAB production, the Actinobacteria were not dominant and did not increase over time suggesting other microorganisms were producing NDAB. 98 In order to test this hypothesis, NDAB production under several experimental conditions (all anoxic systems) was investigated and compared (Table 6.1). NDAB has not been detected in any of the strictly abiotic systems tested. Reduced AQDS
(AH 2QDS), surface bound Fe(II), and sterilized sediment incubations did not generate
NDAB during RDX degradation. NDAB was only detected in experimental bottles with microbial activity, regardless of the presence or absence of electron shuttling compounds.
Previous cell suspensions of Geobacter metallireducens and Shewanella oneidensis also generated NDAB. NDAB was produced (1 5 M) only when cells were present, irrespective of the presence of electron shuttles. Hydroquinones and reduced humic substances did not generate NDAB. However, cell suspensions of an isolate from the sediment used in this study, Desulfotomaculum strain MJ1, did not generate NDAB despite rapid RDX transformation (data not shown). G. metallireducens and S. oneidensis are Fe(III) reducing microorganisms, whereas strain MJ1 is an obligate fermenter. This suggests that Fe(III) reduction may be correlated with biological NDAB formation.
The figure 6.1 data suggest that NDAB was produced during (and likely from)
MNX biodegradation (Figure 6.1B, D, and E). Zhao et al suggested that NDAB formation from RDX through MNX with Clostridium bifermentans HAW 1 was due to abiotic reaction (hydrolysis) of MNX, not biological activity (Zhao et al., 2003b).
We used resting cell suspensions of S. oneidensis to verify whether MNX was further converted to NDAB and if living cells were necessary (Figure 6.3). S. oneidensis was selected because past data with RDX indicate that it produces more
NDAB than G. metallireducens. S. oneidensis produced approximately 6 7 M of
NDAB from MNX degradation regardless of the presence or absence of AQDS or lactate (Figure 6.3A and 6.3B). It is not surprising that S. oneidensis reduced MNX 99 without lactate because the cells utilize decayed cellular biomass as a substrate for respiratory processes which were observed in the previous studies (Fredrickson et al.,
2000a; Kwon and Finneran, 2008b). Abiotic MNX degradation by AH 2QDS produced stoichiometric amount of HCHO (1 MNX 3 HCHO) (Figure 6.3C), but did not produce NDAB (Figure 6.3B). The rate and extent of MNX reduction were faster at the high pH (8.2) than the low pH (7.2) which was consistent with the previous results (Kwon and Finneran, 2008a). The sediment and cell suspension data demonstrate that NDAB production is due to anaerobic, biological RDX transformation, with MNX as a key intermediate.
Additional sediment experiments with NDAB as a starting compound did not show NDAB degradation within 44 days, although approximately 50% of RDX was mineralized into CO 2 under the same experimental condition (data not shown).
Transformation of RDX to NDAB as a dead end product by Rhodococcus sp. DN22 has been proposed previously (Fournier et al., 2002). However, it is also reported that
NDAB degraded to N 2O and CO 2 by soil bacteria Methylobacterium JS178 (Fournier et al., 2005) and by the fungus Phanerochaete chrysosporium (Fournier et al., 2004a).
None of these NDAB degrading organisms were present under the anoxic conditions of this experiment, as none were detected by molecular analyses (though P. chrysosporium detection would have required eukaryotic molecular tools that were not utilized as part of the study. However, no reports exist of anaerobic P. chrysosporium metabolism related to RDX or iron).
Terminal electron accepting processes (TEAPs) correlated with RDX degradation
The geochemical data collected over time demonstrate that RDX reduction was related to Fe(III) reduction rather than nitrate, sulfate reduction, or methanogenesis (Figure 6.5). Nitrate was reduced in all acetate amended incubations 100 (acetate alone, and acetate with AQDS or HS) at least 10 days prior to the onset of
RDX reduction (Figure 6.5A). Nitrate was completely reduced in the sterilized incubations at day 30. It is possible that serial autoclaving does not completely kill all microorganisms in this sediment. RDX was not transformed in sterile bottles. Sulfate was not reduced in any incubation within the time frame of the experiments (Figure
14 6.5B) and sulfide was not detected (data not shown). CH 4 was not produced in any
[14 C] RDX amended incubations (see below). However, reduced iron (% reduced Fe) in AQDS plus acetate incubations rapidly increased to 13% at day 20 and accumulated up to 60% at day 32 (Figure 6.5C). The increase in reduced iron in
AQDS amended incubations corresponds directly to RDX reduction and metabolite
(e.g., nitroso metabolites and ring cleavage products) production. Reduced Fe (%) in
HS amended acetate incubations increased to 4% by day 32, which was slightly greater than acetate alone (at 1% accumulation by day 32). Reduced Fe (%) did not increase in other treatments.
Reduced Fe(II) reacts directly (and sometimes rapidly) with RDX and its nitroso derivatives. While it has not been definitively shown to reduce the ring cleavage metabolites, it is possible based on speculated degradation pathways (Zhao et al., 2003b; Kwon and Finneran, 2008a; Larese Casanova and Scherer, 2008; Naja et al., 2008). It is not surprising then that only AQDS amended incubations had a net increase in reduced Fe(II). The rate of Fe(III) reduction with electron donor alone or purified HS may be equivalent to the rate of Fe(II) mediated RDX (and metabolite) reduction. AQDS, on the other hand, increases the rate of Fe(III) reduction sharply as
3+ electron transfer to Fe(III) from AH 2QDS is very fast (approximately 11 mol Fe reduced/second). Although Fe(III) reduction dominated in all incubations, the net increase in reduced iron was only significant in the AQDS amended bottles. Direct electron transfer from reduced electron shuttles to RDX is also a relevant pathway and 101 some proportion of reducing equivalents from reduced electron shuttles will directly reduce RDX and its metabolites (Kwon and Finneran, 2006, 2008a). The results indicate that Fe(III) reduction is an important TEAP in electron shuttling mediated
RDX remediation in aquifer sediment and this strategy will be successful in both sediments with and without high concentrations of bioavailable iron, the latter being catalyzed by electron shuttling.
[14 C-RDX] Mineralization
Uniformly labeled [ 14 C] RDX was amended at day 32 of the non radiolabeled experiment (not enough material was available for parallel experiments so [ 14 C] RDX was added when MNX was below detection in all acetate amended incubations,
14 which also corresponded to Fe(III) reducing activity). CH 4 was not produced from
14 14 C labeled RDX in any incubation (Figure 6.6). CO 2 production was higher in
14 14 AQDS and HS amended incubations; 40 50% of [ C] RDX was oxidized to CO 2
(Figure 6.6). Acetate alone incubations mineralized 12% of the RDX. RDX mineralization began without a lag period, indicating there was no acclimation period for the active microorganisms, which at the time was dominated by Fe(III) reducers.
The increased mineralization with electron shuttling compounds was likely due to stimulation of the most appropriate community for further oxidation of ring cleavage compounds. HCHO was the dominant reaction product generated by AH 2QDS
(Figure 6.3C). HCHO can be oxidized by Fe(III) reducing microorganisms (Shrout et al., 2005), and is the most likely mineralization precursor in these experiments
Temporal variation of microbial communities
Clone libraries were assembled with 16S rRNA genes amplified from RDX contaminated aquifer material at key time points relative to RDX degradation and 102 geochemical changes. Microorganisms closely related to known Fe(III) reducers and affiliated with in situ Fe(III) and U(VI) reduction predominated after RDX reduction and immediately prior to mineralization (Figure 6.7). The most interesting characteristic of the Fe(III) and electron shuttle reducing community is that it was not dominated by δ proteobacteria or members of the Geobacteraceae (they were present but not dominant). This demonstrates that Fe(III) reducing microbial communities other than Geobacteraceae can be stimulated in situ and catalyze the primary reactions of interest.
Initial (time = 0) 16S rRNA gene sequences in the aquifer material were highly similar to α proteobacteria (Genus Ochrobactrum ) and predominated in all incubations. However, this class decreased over time; 60 to 24% in AQDS plus acetate incubations, 82 to 46% in HS plus acetate incubations, and 78 to 50% in acetate alone incubations. Most of the clones identified in the current study were similar to Ochrobactrum anthropi strain CCUG 43892 with more than 99% similarity
(Table 6.2). However, Ochrobactrum anthropi strain YZ 1 is not reported to reduce hydrous Fe(III) oxide with acetate (Zuo et al., 2008). In contrast to the sequences at day 0, 16S rRNA gene sequences at day 32 were highly similar to known Fe(III) reducing microorganisms. AQDS amended microbial communities included
Pseudomonas and Geobacteraceae that increased from 29% to 67%. In the presence of HS, 52% of sequences at day 32 were idenfied as β proteobacteria (families
Oxalobacteraceae ). Although Oxalobacteraceae isolates are not known Fe(III) reducers, phylotypes similar to these have been recovered in uranium contaminated material during stimulated Fe(III) reduction (Reardon et al., 2004; Akob et al., 2007;
Akob et al., 2008; Li and Krumholz, 2008). Nitrate was depleted by day 6; therefore, it could not be the terminal electron acceptor promoting microbial activity at day 32.
Sequences in acetate alone incubations were 30% related to β proteobacteria 103 (families Oxalobacteraceae and Comamonadaceae ) and 8% related to γ proteobacteria (Genus Pseudomonas ).
Pseudomonas and Geobacteraceae in AQDS plus acetate incubations were not unexpected because both genera contain well known dissimilatory Fe(III) reducers
(Lovley, 1997; Lu et al., 2002). However, previous reports would have us predict more presence of phylotypes within the Geobacteraceae (Snoeyenbos West et al.,
2000; Holmes et al., 2002). These data suggest that it is unlikely a single group of organisms can become dominant in every environment represented by a particular metabolic process. In this case the β proteobacteria were dominant phylotypes relative to the δ proteobacteria .
β proteobacteria have been enriched in alternate Fe(III) reducing conditions such as in ethanol and glucose amended nitrate reducing and Fe(III) reducing MPN cultures (Akob et al., 2008), in uranium contaminated subsurface sediments (Akob et al., 2007), in enrichment culture of freshwater wetland sediment microorganisms
(Weber et al., 2006), and in a specular hematite (Fe 2O3) medium (Reardon et al.,
2004). In particular, Weber et al (Weber et al., 2006) showed that β proteobacteria increase from 3% to 60% when experimental condition shifted from nirate reducing to Fe(III) reducing conditions. Reardon et al (Reardon et al., 2004) also reported that the majority of hematite associated community formed in the pristine area was affiliated with β proteobacteria , and one of the clones was designated as S E105, which was 98% similar to clones PTA 29 and PTA 31 in the current study (Table 6.2).
β proteobacteria were also enriched in Tc(VII) reducing sediments in which Tc(VII) was reduced concurrently with Fe(III) after nitrate was depleted (Li and Krumholz,
2008). Li et al (Li and Krumholz, 2008) described several clones (M0C22 and
M20C4) that proliferated during metal reduction, and both were 98% similar to clones
PTA 29 and PTA 31 in the current study (Table 6.2). The clones identified in this 104 study were enriched during Fe(III) reduction (long after nitrate reduction and prior to sulfate reduction). Clone PTA 29 was also 97% similar to Herbaspirillum sp. K1
(family Oxalobacteraceae ) which is a facultative bacterium isolated from contaminated groundwater that degrades tetrachlorophenol (Männistö et al., 2001).
Different microbial communities were enriched when AQDS versus HS was used (Figure 6.7). At day 20 and 32, microbial communities in AQDS amended incubations were predominantly γ proteobacteria (Genus Pseudomonas ), while those in HS amended incubations were β proteobacteria (families Oxalobacteraceae and
Comamonadaceae ). Since many Fe(III) reducing microorganisms can also reduce humic substances (Coates et al., 1998; Lovley et al., 1998; Kwon and Finneran,
2008b), γ proteobacteria and β proteobacteria enriched from AQDS and HS amended incubations, respectively, may utilize both electron shuttling compounds and
Fe(III) (directly) as electron acceptors. However, the differences of molecular structure between AQDS (e.g., low molecular weight, simple functional group) and humics (e.g., high molecular weight and multiple functional group) enriched different microorganisms.
Environmental Relevance
Several unique RDX degradation patterns were identified by these data. First,
RDX is most completely mineralized (To CO 2) in the presence of electron shuttling compounds, due to increased Fe(III) reducing microbial activity. Second, NDAB was produced directly by Fe(III) reducing microorganisms under anoxic conditions.
Previously, anoxic NDAB production was reported as abiotic. This is important because NDAB accumulation may correspond to an increase in Fe(III) reduction mediated RDX transformation, which will ultimately lead to more mineralization.
RDX reduction correlated to Fe(III) reduction rather than nitrate or sulfate reduction. 105 Finally, the Fe(III) reducing microorganisms enriched in situ were outside of the
“typical” group ( Geobacteraceae ) associated with acetate + electron shuttle amendment, demonstrating that RDX contamination exerts a selective pressure on the microbial community.
The data suggest that RDX degrades to a variety of intermediates but is ultimately mineralized more quickly and completely with electron shuttling compounds. Adding AQDS or HS prevented accumulation of the nitroso metabolites relative to electron donor alone; substrate addition has been the primary strategy for
RDX to date. The ring cleavage metabolite NDAB increased with electron shuttles but not in other incubations, indicating that NDAB could be produced with the microorganisms utilizing electron shuttling compounds + Fe(III) under anoxic conditions. All reactions were catalyzed by native microorganisms, in particular
Fe(III) reducing bacteria.
Targeting Fe(III) and/or electron shuttle reduction opens up a major group of indigenous microorganisms to stimulate these reactions. Fe(III) and humic reducing microorganisms are ubiquitous (Coates et al., 1998); therefore, the strategy using electron shuttles and Fe(III) and humic reducing microorganisms will likely work at many sites. Finally, Fe(III) reduction in this study was driven by organisms primarily within the β and γ proteobacteria . This is vastly different from past reports which suggested that δ proteobacteria are most dominant during Fe(III) reduction, and we intend to follow up by investigating Fe(III) and electron shuttle reduction within genera that were not previously tested for Fe(III)/shuttle reduction, but which clearly are important to this metabolism in situ.
106 60
50
40 AQDS+acetate M) AQDS-acetate 30 HS+acetate HS-acetate RDX ( RDX 20 active unamended 10 active+acetate A sterile+acetate 0
3 M) 2
MNX MNX ( 1 B 0 10
8 M) 6
4 DNX ( DNX 2 C 0 40 D
30 M) 20 TNX ( TNX
10
0 12 E 10
M) 8
6
NDAB ( NDAB 4
2
0 0 10 20 30 time (d) Figure 6.1. RDX reduction (1A) and the production and loss of MNX (1B), DNX (1C), TNX (1D), and NDAB (1E) in contaminated aquifer material from the Picatinny Arsenal that has been amended with 10mM acetate plus the electron shuttles AQDS (100 M) or purified humic substances (HS) (0.15 g/L), or which had only been amended with the electron donor acetate (10 mM). The aquifer material was aerated before beginning the experiments. Results are the means of triplicate analyses and bars indicate one standard deviation
107 40
AQDS+acetate AQDS-acetate A 30 HS+acetate HS-acetate
M) active unamended active+acetate 20 sterile+acetate HCHO ( HCHO
10
0
6 B M) 4
MEDINA ( MEDINA 2
0 50 C
40 M) 30 ( + 4
NH 20
10
0 0 10 20 30 time (d) Figure 6.2. Production of RDX ring cleavage products, HCHO (A), MEDINA (B), + and NH 4 (C), in contaminated aquifer material from the Picatinny Arsenal that has been amended with the electron shuttles AQDS (100 M) or purified HS (0.15 g/L), or which has only been amended with the electron donor acetate (10 mM). The aquifer material was aerated before beginning the experiments. Results are the means of triplicate analyses and bars indicate one standard deviation.
108 40
30 M) 20 MNX (
10 A 0
10 MR1+AQDS+Lactate MR1+Lactate MR1 alone M)
8 AH 2QDS ) ( 3 AH 2QDS @pH 8.2 O 3 AQDS alone N 6 5 Lactate alone H 2 unamended 4 NDAB (C 2 B
0 120
100
80 M)
60 HCHO ( HCHO 40
20 C
0 0 5 10 15 20 25 30 time (h) Figure 6.3. MNX reduction (A) and the production of NDAB (B) and HCHO (C) in the resting cell suspensions of Shewanealla oneidensis MR1 that has been amended with the electron shuttles AQDS (100 M), or which has only been amended with the electron donor lactate (10 mM). Control experiments were performed with AH 2QDS at pH 7.2 and 8.2 or with AQDS alone or lactate alone. Results are the means of triplicate analyses and bars indicate one standard deviation
109 Figure 6.4. Probable RDX degradation routes based on products identified and reported degradation pathways in the sediment incubations. Compounds in square brackets were not determined; [C] represents unidentified carbon intermediates. A = abiotic pathway(s), B = biological pathway(s). The > and < indicate whether the products were more significant in the abiotic or biological pathway (based on the mass balance). RDX degradation pathway in the sediment incubations is considerably different from our previous study (Kwon and Finneran, 2008a) because the past experiments were performed with G. metallireducens GS 15 as a model electron shuttle reducing microorganism, which was not the dominant Fe(III) reducer present in the current study (Table 6.2). GS 15 generated a small amount of MNX; DNX and TNX were insignificant. However, significant amount of nitroso metabolites accumulated in sediment incubations. Another difference between two studies is that AQDS amended GS 15 cell suspensions rapidly transformed RDX to HCHO (ca. 50% of the carbon mass balance) and only 7 20% of RDX was mineralized (per the carbon mass balance). Here, HCHO was limited (< 6% of the carbon mass balance) and mineralization to CO 2 was significant (~50% of the carbon mass balance) indicating that HCHO oxidation may be prevalent within the Fe(III) reducing microbial community enriched in situ (see below) (Figure 6.7). In addition, NDAB is generated most likely from MNX degradation based on the sediment and pure cell incubations (Figure 6.1 and 6.3) (Refer to the results and discussion section in the main text).
110 0.6 Onset of RDX reduction 0.5 A 0.4
0.3
Nitrate (mM) 0.2
0.1
0.0
2.0 B
1.5
1.0 Sulfate (mM) Sulfate
0.5
0.0
AQDS+acetate 60 AQDS-acetate HS+acetate HS-acetate active unamended 40 active+acetate sterile+acetate
reduced(%) Fe 20 C
0 0 10 20 30 time (d) Figure 6.5. Terminal electron accepting processes (TEAPs) correlated with RDX degradation. Reduction of nitrate (A), sulfate (B), and iron (C) as TEAPs in contaminated aquifer material from the Picatinny Arsenal that has been amended with 10 mM acetate plus the electron shuttles AQDS (100 M) or purified humic substances (HS) (0.15 g/L), or which had only been amended with the electron donor acetate (10 mM). The aquifer material was aerated before beginning the experiments. The aquifer material was amended with 0.5 mM nitrate and 1.5 mM sulfate before beginning the experiments. Results are the means of triplicate analyses and bars indicate one standard deviation.
111
80 AQDS+acetate AQDS-acetate HS+acetate
(%) 60
2 HS-acetate active unamended active+acetate sterile+acetate 40
20 mineralization to to CO mineralization
0 0 2 4 6 8
time (d) Figure 6.6. RDX mineralization in contaminated aquifer material from the Picatinny Arsenal. U [14 C] RDX was amended at day 32 of non radiolabeled RDX experiment. No additional amendments were added or re added at this time point. Results are the means of triplicate analyses and bars indicate one standard deviation
112
Day 0 Day 20 Day 34
AQDS+ acetate
HS+ acetate
Bacteroidetes Verrucomicrobia Actinobacteria Chloroflexi Firmicutes Spirochaetes α-proteobacteria β-proteobacteria γ-proteobacteria acetate δ-proteobacteria alone ε-proteobacteria unidentified
Figure 6.7. Temporal variation of microbial community composition in contaminated aquifer material from the Picatinny Arsenal that has been amended with 10 mM acetate plus the electron shuttles AQDS (100 M) or purified humic substances (HS) (0.15 g/L), or which had only been amended with the electron donor acetate (10 mM)
113
* * *
Figure 6.8. Phylogenetic tree of non proteobacteria (A) and proteobacteria (B) based on 16S rRNA gene sequence cloned from aquifer sediment incubations. Branch points were supported by Jukes Cantor distance and maximum likelihood methods. (*) indicates known Fe(III) reducing microorganisms. The scale bar indicates 0.2 (A) and 0.08 (B) change per nucleotide position. 114
Figure 6.8. (continued) Phylogenetic tree of non proteobacteria (A) and proteobacteria (B) based on 16S rRNA gene sequence cloned from aquifer sediment incubations. Branch points were supported by Jukes Cantor distance and maximum likelihood methods. (*) indicates known Fe(III) reducing microorganisms. The scale bar indicates 0.2 (A) and 0.08 (B) change per nucleotide position.
115 e
h t r e o t t c t a r o b e f n o e e = R e G ( A h . t = 6, 2008b) indicate lactate; GS15 species strain MJ1; N Desulfotomaculum strain MR1; MJ1= experimental conditions. The numbers in parenthesis d from our previous studies (Kwon and Finneran, 200 rly crystalline Fe(III) hydroxide; Ac= acetate; Lc= Shewanella oneidensis Shewanella strain GS15; MR1= Table 6.1. Summary of NDAB production from various experimental pH. Fe(II)= aqueous Fe(II); FeGel= poo metallireducens applicable. The data of GS15 and MR1 were reproduce results and discussion section in the main text). 116 Table 6.2. Summary of phylogenetic affiliation and distribution of 16S rDNA clones from clone libraries of sediment samples from contaminated aquifer material from the Picatinny Arsenal. The actual number of representative clones at day 0, 20, 32 and the nearest relatives to each clone were shown. *Clones showed 98% similarity to clone S E105 identified by (Reardon et al., 2004) and clone M0C22 and M20C4 identified by (Li and Krumholz, 2008) (Refer to the results and discussion section).
No. of related clones in each incubation with time Phylogenetic group Clone ID Nearest relative (GenBank accession no.) % similarity AQDS+acetate HS+acetate acetate alone 0d 20d 32d 0d 20d 32d 0d 20d 32d PTA-01 Winogradskyella sp. MOLA 345 (AM945575) 99 1 Bacteroidetes PTA-02 Bacteroidetes bacterium clone: Y4 (AB116461) 97 1 PTA-03 Gelidibacter sp. IMCC1914 (EF108219) 99 1 Verrucomicrobia PTA-04 Opitutus sp. VeSm13 (X99392) 97 1 PTA-05 Nocardioides sp. V4.BE.17 (AJ244657) 99 1 Actinobacteria PTA-06 Arthrobacter T. 70 (AJ864860) 100 1 PTA-07 Actinobacterium clone 3G1820-35 (DQ431885) 100 4 5 PTA-08 Chloroflexi bacterium clone XME56 (EF061972) 98 1 Chloroflexi PTA-09 Chloroflexi bacterium clone MSB-5bx1 (DQ811889) 96 1 PTA-10 Firmicutes bacterium clone GASP-KC1W2_A08 (EU299417) 99 1 Firmicutes PTA-11 Clostridium thermocellum DSM 1237 (L09173) 96 1 PTA-12 Desulfitobacterium chlororespirans (U68528) 98 1 PTA-13 Spirochaete clone SIMO-2393 (AY711759.1) 99 1 1 Spirochaetes PTA-14 Spirochaetes bacterium SA-10 (AY695841) 97 1 PTA-15 Ochrobactrum anthropi strain CCUG 43892 (AM490611) 100 14 15 26 20 9 23 13 7 PTA-16 Ochrobactrum anthropi strain CCUG 43892 (AM490611) 99 12 12 6 12 19 13 14 23 15 PTA-17 Sphingomonas aerolata strain NW12 (AJ429240) 99 5 6 2 1 2 PTA-19 Rhodoplanes sp. clone GASP-KC3W2_B10 (EU300474) 100 1 Alphaproteobacteria PTA-20 Mesorhizobium sp. CCBAU 65327 (EU074203) 99 7 1 1 PTA-22 Anderseniella baltica strain BA141T (AM712634) 98 1 PTA-23 Alpha proteobacterium clone lhap10 (DQ648968) 100 1 PTA-24 Alpha proteobacterium clone lhap10 (DQ648968) 99 1 PTA-25 Dechloromonas denitrificans strain ED1T (AJ318917) 100 1 PTA-26 Delftia acidovorans strain G-1-15 (EF102823) 100 1 2 PTA-27 Ramlibacter sp. clone GASP-MA4S3_C01 (EF664127) 99 1 Betaproteobacteria PTA-28 Ramlibacter P-8 (AM411936) 98 1 PTA-29* Herbaspirillum seropedicae strain BA153 (AF164062) 98 12 8 17 2 8 13 PTA-31* Aquaspirillum autotrophicum (AB074524) 99 7 1 PTA-32 Gamma proteobacterium clone MVP-73 (DQ676357) 98 1 PTA-33 Pseudomonas fluorescens strain 29L (EF424136) 99 14 32 4 1 1 8 Gammaproteobacteria PTA-34 Pseudomonas stutzeri strain HS-D36 (EF515811) 99 4 PTA-35 Pseudomonas fluorescens strain 29L (EF424136) 98 1 PTA-36 Geobacter sp. clone: GH-11 (AB293318) 98 1 2 2 Deltaproteobacteria PTA-37 Geobacter argillaceus strain G12 (DQ145534) 96 1 Epsilonproteobacteria PTA-38 Thiomicrospira sp. CVO (U46506) 98 1 Unidentified 3 1 2 1 4 4 2 total 52 56 55 50 52 50 50 54 50
117 CHAPTER 7
CHARACTERIZATION OF A HEXAHYDRO 1,3,5 TRINITRO 1,3,5
TRIAZINE (RDX) DEGRADING, NOVEL ANAEROBE ISOLATED FROM
CONTAMINATED AQUIFER MATERIAL
SUMMARY
A Gram stain negative, fermentative, non spore forming, motile without flagella, rod shaped bacterium, designated strain MJ1 T, was isolated from an RDX contaminated aquifer. On the basis of 16S rRNA gene sequencing, strain MJ1 is relegated to Firmicutes . The DNA G+C content was 42.8 mol %. Fermentative growth occurred on glucose and citrate in defined basal medium. The bacterium is a strict anaerobe that grows between at pH 6.0 and pH 8.0 and 18 °C and 37 °C. The culture did not grow with hexahydro 1,3,5 trinitro 1,3,5 triazine (RDX) as the electron acceptor or mineralize RDX under these conditions. However, MJ1 transformed RDX into MNX, methylenedinitramine (MEDINA), formaldehyde, formate, ammonium, nitrous oxide, and nitrate. The nearest phylogenetic relatives
(95% similarity) with valid taxonomic designations were Desulfotomaculum guttoideum . However, MJ1 also showed close similarity with Clostridium celerecrescens DSM 5628 (95%), Clostridium indolis DSM 755 (94%) and
Clostridium sphenoides DSM 632 (94%). On the basis of morphological, physiological, and genetic information, MJ1 is a new species in the family
Clostridiaceae .
118
INTRODUCTION
Several Clostridium species (e.g., C. acetobutylicum , C. bifermentans , C. sp.
EDB2) have been reported to degrade the nitramine explosive compound hexahydro
1,3,5 trinitro 1,3,5 triazine (RDX) (Regan and Crawford, 1994; Zhang and Hughes,
2003; Zhao et al., 2003b; Bhushan et al., 2004; Bhushan et al., 2006). RDX is a contaminant of concern at many military sites and live fire training installations
(Spalding and Fulton, 1988; Haas et al., 1990; Hawari and Halasz, 2002).
Groundwater contaminated by RDX is of growing environmental concern because of its associated human health effects (U.S., 2004). In situ bioremediation relying on
Clostridium species has been proposed as a cost effective and eco friendly strategy for eliminating RDX from subsurface environments (Zhao et al., 2003a). The genus
Clostridium is comprised of Gram positive, obligate anaerobes capable of producing endospores. They represent one of the largest genera of the prokaryotes (>170 species) commonly present in subsurface system and have a wide range of metabolic properties (Collins et al., 1994; Hippe et al., 2003). The study presented here investigates an isolate, MJ1, which is closely related to Clostridium . MJ1 degraded
RDX as effectively as the previously reported Clostridium species.
MATERIALS AND METHODS
Strain MJ1 was originally enriched and isolated from the RDX contaminated aquifer material used in the chapter 6. Anaerobic agar slant solidified with the modified Wolfe’s fresh water medium were used for the isolation. A single colony was selected and transferred to the fresh water medium with 20 mM of lactate and 50 mM of Fe(III) citrate as the sole electron donor. The basal anaerobic medium
1 consisted of (g l unless specified otherwise): NaHCO 3 (2.5), NH 4Cl (0.25), 119 NaH 2PO 4 H 2O (0.6), KCl (0.1), modified Wolfe’s vitamin and mineral mixtures (each
1 10 ml l ) (Lovley et al., 1993) and 1 mL of 1 mM Na 2SeO 4. The medium was buffered using a 30mM bicarbonate buffer equilibrated with 80:20 (vol/vol) N 2:CO 2, as previously described (Finneran et al., 2003). Standard anaerobic and aseptic culturing techniques were used throughout (Lovley and Phillips, 1988). The medium was sparged with anaerobic gases passed over a heated, reduced copper column to remove trace oxygen from the gas line. Culture tube/bottle headspaces were flushed with the same gas mixture and sealed under an anaerobic headspace using a thick butyl rubber stopper fastened with an aluminum crimp. Cultures were maintained in
160mL anaerobic serum bottles; individual experiments utilized different culture tubes or bottles, as described below. Culture purity was checked by microscopic examination.
In order to investigate kinetics of RDX degradation and metabolite production, the resting cell suspension of MJ1 was prepared as described previously (Kwon and
Finneran, 2006). MJ1 was grown anaerobically in freshwater medium with glucose
(10mM). Experimental tubes were prepared by sparging approximately 5.0ml of
30mM bicarbonate buffer with anaerobic N 2:CO 2 (= 80 : 20; pH = 6.8) and sealing the buffer under an anaerobic headspace. All chemicals were amended from concentrated, anaerobic, sterile stock solutions. RDX was amended at the concentration of 60 M in the resting cell suspension of MJ1. An aliquot (0.3 ml) of the resting cells was added to the sealed pressure tubes to initiate each experiment.
Individual tubes were incubated at 30°C. Samples (0.5 ml) were collected periodically via anaerobic syringe and needle. Analytical methods for RDX and its metabolites were described in Appendix.
Gram staining was performed according to the standard procedure (Bergey and Holt, 2000). Cell morphology and motility were investigated with phase contrast 120 microscope (Zeiss Axioskop, Zeiss Inc.) and Scanning Electron Microscope
(SEM)(JEOL JSM 6060LV) at 20kV with cells harvested from a log phase culture.
SEM photos were produced by The Frederick Seitz Materials Research Laboratory at the University of Illinois. G+C content was performed by DSMZ characterization services using an HPLC method.
Microbial growth was quantified by inoculating 0.5 ml of glucose grown MJ1 into 9.5 ml fresh water media with 10 mM of glucose or 50 mM of citrate. MJ1 growth was tested at various temperatures (4, 18, 30, 37 and 60 °C) and pH values
(5.0 9.0 in 1.0 pH unit increments) in fresh water media with glucose or citrate.
Penicillin, streptomycin and tetracycline (100 mg l 1) were applied to determine if they inhibited the growth of MJ1 or if the culture was resistant to commonly used antibiotics. Growth tests were performed either in triplicate (for temperature values, pH values and antibiotics tests) or in single tube (for alternative electron donors and acceptors) for 48h to 2 weeks. Growth (in any experiment) was monitored by measuring the turbidity (optical density) by spectrophotometry at 600nm
(GENESYS TM 2, Thermo Spectronic Inc.).
The nearly complete 16S rRNA gene of strain MJ1 was used to establish phylogenetic placement. DNA was extracted using the Fast DNA extraction kit (MP
Biomedicals, LLC) with a bead beating apparatus according to the manufacturer’s instructions. The 16S rRNA gene was amplified with primers Eub27F and Eub1492R with an initial denaturation step at 94°C for 4 min, followed by 35 cycles of 94°C
(30s), 50°C (30s), and 72°C (1 min 30s) with a final extension at 72°C for 7 min, and holding at 4 °C. The sequence was purified by QiaQuick PCR purification kit
(QIAGEN) and sequenced at the University of Illinois Core Sequencing Center. The
GenBank Accession Number for the MJ1 full 16S rRNA gene sequence is FJ008042.
121 The initial percent of identity was determined using the National Center for
Biotechnology Information (NCBI) Basic local alignment search tool (BLAST), which compares the sequence to the Genbank database. Phylogenetic trees were constructed using the program Geneious 3.7.0 based on the maximum likelihood method with 100 bootstrap replicates. Reference microorganisms selected were closely related Desulfotomaculum and Clostridium species.
RESULTS AND DISCUSSION
The bacterium formed transluscent, convex, circular colonies on LB media plates. Cells stained gram negative and were straight or slightly curved rods. The cells were approximately 3 5 m long and 0.6 0.7 m in diameter (Figure 7.1). MJ1 was motile, though flagella were not apparent in any preparation. Spores were not detected either by heat shock at 82 °C for 15 minutes or aeration for 10 days, as confirmed by microscopic examination. MJ1 stained negatives. Several
Desulfotomaculum showed negative response for gram staining. However, the typical structure of gram positive bacteria in their cell walls reported previously (Nazina and
Pivovarova, 1979). In most case, Clostridium showed positive response for gram staining. This led the possibilities that MJ1 has the typical structure of gram positive, but stains negative. The G+C content was 42.8 mol %. Although Clostridium species generally have relatively low GC content, they show a wide range of GC contents
(22 55 mol%) (Collins et al., 1994; Hippe et al., 2003).
MJ1 grew in the fresh water medium with glucose (10mM). MJ1 grew well between 18 °C and 37 °C; 30 °C was the optimal growth temperature. Growth at 37 °C and 18 °C was relatively fast but the extent of growth was low. The culture did not grow at 4 °C or 60 °C (Figure 7.2A). MJ1 grew best at pH 7.0. However, growth at pH
6.0 and pH 8.0 was also relatively fast and the growth kinetics were similar (Figure 122 7.2B). The order of the growth at different pH values was pH 7.0 > pH 8.0 > pH 6.0.
The culture did not grow at pH 5.0 and 9.0. MJ1 also grew in the fresh water medium with citrate (50 mM) and showed a similar pattern. However, the growth rate with citrate was much slower than with glucose and MJ1 grew at pH 5.0 with citrate (data not shown). Growth was completely inhibited by 100 mg l 1 streptomycin and tetracycline. MJ1 had a 69 hour lag phase with 100 mg l 1 penicillin prior to the onset of growth.
The fermentation products of strain MJ1 from glucose (10 mM) were hydrogen (0.74 mM), acetate (9.3 mM), and ethanol (1 mM). No or little lactate, propionate, citrate, butanol, or butyrate was detected. Growth was not supported in fresh water medium with acetate (20 mM) or lactate (20 mM) as the sole electon donor and nitrate (2 mM), nitrite (1 mM), sulfate (5 mM), thiosulfate (10 mM), or sulfite (10 mM) as the sole electron acceptor. Since MJ1 did not reduce sulfates, sulfites, and other sulfur compounds, it is difficult to classify MJ1 as a
Desulfotomaculum species.
The nearest phylogenetic relative for which a cultured isolate is available was
Desulfotomaculum guttoideum (DSM 4024) with 95% identity. Other related
Desulfotomaculum species included Desulfotomaculum geothermicum DSM 4042
(ATCC 49053) with approximately 94% identity (Figure 7.3). MJ1 also showed close similarity with Clostridium celerecrescens DSM 5628 (95%), Clostridium indolis
DSM 755 (94%) and Clostridium sphenoides DSM 632 (94%). In fact, it is reported that Desulfotomaculum guttoideum might be a misidentified species because it shares extremely high 16S rDNA similarity with several Clostridium species such as C. sphenoides and C. celerecrescens (Stackenbrandt et al., 1997). Therefore, it may be more reasonable to classify MJ1 as Clostridium species rather than Desulfotomaculum
123 species. Table 7.1 summarizes the physiological similarities and differences of MJ1 amongst Clostridium species which are able to transform RDX.
A resting cell suspension of MJ1 completely reduced RDX within 20 hours and produced several metabolites including MNX, methylenedinitramine (MEDINA), formaldehyde, formate, ammonium, nitrous oxide, and nitrate (Figure 7.4). However,
MJ1 did not mineralize it to CO 2 (data not shown) while GS15 mineralized approximately 10% of RDX to CO 2. Although the rate and extent of RDX degradation by MJ1 was similar to those by Geobacter metallireducens GS15, product distribution was very different, suggesting MJ1 metabolic activities are different from GS15. MJ1 was also compared to several RDX degrading Clostridium species. Table 7.2 summarizes the similarities and differences amongst these strains.
RDX degradation products by MJ1 are quite similar to those generated by Clostridium sp. EDB2.
In summary, MJ1 is an RDX reducing microorganism that uses glucose as the sole carbon and energy source. MJ1 has only 95% similarity to the most closely related Desulfotomaculum guttoideum or other Clostridium species. In addition, alternate physiological characteristics of MJ1 warrant that it is a new species.
Therefore, this isolate is proposed as a new species within the Genus Clostridium .
Strain MJ1 is the type strain and has been submitted to the American Type Culture
Collection (ATCC) and the Deutsche Sammlung von Mikroorganismen (DSMZ).
DESCRIPTION OF A NOVEL SPECIES MJ1
Cells are motile, gram negative (by staining) rods of approximately 3 5 m in length by 0.6 0.7 m in diameter. The cells do not form terminal spores after heat treatment. Flagella were absent. Colonies on solid medium are smooth, clear, rounded, and convex with a diameter of approximately 1mm. The cells grow well in 124 minimal freshwater medium with glucose or citrate under anaerobic conditions. The culture is a strict anaerobe that reduces RDX. Growth with the primary carbon substrate, glucose, is faster than with citrate. Fumarate, nitrate, nitrite, sufate, sulfite,
Fe(III) EDTA and Fe(III) NTA were not used as electron acceptors. The optimal growth temperature was 30 °C, but the culture grew well between 18 °C and 37 °C.
The optimal growth pH was 7.0, but the cells grew well from 6.0 to 8.0. The G+C content was 42.8 mol %. This culture was originally isolated from explosive contaminated aquifer material.
125
Figure 7.1. Scanning electron microscopy (SEM) image of MJ1. Cells were approximately 3 5 m long and 0.6 0.7 m in diameter. No flagella were identified. The bar is 1 m.
126
1.6
o 1.4 A 4 C 18 oC 1.2 30 oC 37 oC 1.0 60 oC
600 0.8 OD 0.6
0.4
0.2
0.0 1.6 B pH 5 1.4 pH 6 pH 7 1.2 pH 8 pH 9 1.0 600 0.8 OD 0.6
0.4
0.2
0.0 0 20 40 60 80 100 120
time (h) Figure 7.2. Growth curves of MJ1 with glucose (10mM) at various temperature (A) and pH (B). Results are the means of triplicate analyses.
127 Figure 7.3. Phylogenetic position of strain MJ1. The tree is based on the full 16s rRNA gene sequence of MJ1. The sequence was compared to related Desulfotomaculum and Clostridium . Sequences were aligned with the program Geneious 3.7.0. Branch points were supported by maximum likelihood using PHYML and neighbor joining method (Tamura Nei distances (Tamura and Nei, 1993)). Bootstrap values at major nodes are the result of 100 replicates. The bar indicates phylogenetic distance per 100 bases. (*) indicates strains capable of reducing RDX.
128 RDX 100 MNX MEDINA 300 nitrite 80 CHOO - M) N2O + NH 4 200 M) 60 ( + 4
40 NH 100 Concentration ( Concentration 20 A 0 0
100 RDX MNX 300 MEDINA nitrite 80 HCHO M) N2O + NH 4 M)
200 60 ( + 4
40 NH 100 Concentration ( Concentration 20 B 0 0 0 5 10 15 20 25
t (h) Figure 7.4. RDX reduction and metabolite production in the resting cell suspensions of MJ1 (A) and G. metallireducens GS15 (B). MJ1 was incubated without electron donor while GS15 was incubated with acetate (20mM). Results are the means of triplicate analyses. Bars indicate one standard deviation.
129 Table 7.1. Comparison of MJ1 with selected Clostridium species capable of degrading RDX including Clostridium sp. EDB2 (Bhushan et al., 2004; Bhushan et al., 2006), Clostridium bifermentans sp. HAW 1 (Zhao et al., 2003b; Zhao et al., 2004a), Clostridium bifermentans sp. KMR 1 (Regan and Crawford, 1994), and Clostridium acetobutylicum (Zhang and Hughes, 2003) Clostridium sp. C. bifermentans C. bifermentans C. acetobutylicum Characteristics MJ1 EDB2 HAW 1 KMR 1 Shape rod rod rod rod rod Motility + + NA + NA Flagella None Identified many NA NA NA Gram stain variable + + + Spore formation + + + Optimum temperature (°C) 30 30 NA NA NA Temperature range (°C) 18 37 NA 2) NA NA NA Optimum pH 7.0 7.0 NA NA NA pH range 6.0 9.0 NA NA NA NA G+C (mol %) 42.8 53.6 NA NA 30.9 catalase ND 1) NA oxidase ND NA NA NA 1) not determined 2) not available
130 Table 7.2. Comparison of RDX reduction and metabolite distribution by MJ1 with selected Clostridium species including Clostridium sp. EDB2 (Bhushan et al., 2004; Bhushan et al., 2006), Clostridium bifermentans sp. HAW 1 (Zhao et al., 2003b; Zhao et al., 2004a), Clostridium bifermentans sp. KMR 1 (Regan and Crawford, 1994), and Clostridium acetobutylicum (Zhang and Hughes, 2003)
1) Clostridium sp. C. bifermentans C. bifermentans C. acetobutylicum MJ1 2) 3) 4) 5) Compounds EDB2 HAW 1 KMR 1 % N % C % N % C % N % C recovery recovery recovery recovery recovery recovery RDX 1.9 1.9 0.0 0.0 0.0 0.0 0 0 MNX 0.8 0.8 trace trace 0.0 0.0 ND trace DNX trace trace trace trace 0.0 0.0 ND ND TNX trace trace trace trace 0.0 0.0 ND ND MEDINA 11.2 5.6 13.3 6.7 NT NT ND ND NDAB NT 7) NT NT NT NT NT ND ND HCHO NA 8) 1.0 NA 60.0 NA 7.4 ND ND Formic acid NA 46.8 NA 20.0 ND ND ND ND NO 2 11.3 NA 11.7 NA ND NA ND ND + 6) NH 4 53.7 NA ND NA ND* NA ND ND N2O 18.4 NA 46.7 NA 29.5 NA ND ND MeOH NA ND NA ND NA 23 ND ND
CO 2 (mineralization) NA NT NA ND NA 3 ND ND Total mass balance 97.4 56.1 81.7 86.7 29.5 33.4 NA NA 1) mass balance determined after 12 hour reaction 2) mass balance determined after 4 hour reaction 3) mass balance determined after 150 hour reaction which was recalculated from Zhao et al. 2003 4) no metabolites information 5) hydroxylamino/ amino compounds were detected 6) ND: not determined; ND* by interfence 7) NT: not detected 8) NA: not applicable
131 CHAPTER 8 RAW HUMICS EXTRACT AND MILITARY SMOKE DYE AS
EXTRACELLULAR ELECTRON SHUTTLES FOR BIODEGRADATION OF
HEXAHYDRO 1,3,5 TRINITRO 1,3,5 TRIAZINE (RDX)
INTRODUCTION
Extracellular electron shuttles are chemical compounds that stimulate the biodegradation of contaminants by facilitating the transfer of electrons from microorganisms to contaminants (Lovley et al., 1996b; Lovley et al., 1998). The electron shuttles can be used over and over again since they are not consumed in the overall chemical reaction. In general, electron shuttles contain quinone groups, which can be reduced or oxidized and therefore can shuttle electrons from electron donor to electron acceptors such as Fe(III) (hydr)oxides. This can eliminate the need for cell oxide contact by mediating electrons between microorganisms and solid phase minerals (Lovley et al., 1996b). In addition, electron shuttles are believed to stimulate the transformation of Fe(III) (hydr)oxides to more reactive minerals such as magnetite and green rust, which also abiotically reduced organic and inorganic contaminants
(O'Loughlin et al., 2003; Gregory et al., 2004).
Electron shuttle mediated contaminant transformation has been demonstrated for the BTEX compounds (Lovley et al., 1994; Cervantes et al., 2001; Finneran and
Lovley, 2001), methyl tert butyl ether (MTBE) (Finneran and Lovley, 2001; Finneran et al., 2001), carbon tetrachloride (Cervantes et al., 2004; Doong and Chiang, 2005), chlorinated compounds such as dicholoroethene and vinyl chloride (Bradley et al.,
1998), and metals such as uranium (Fredrickson et al., 2000a; Fredrickson et al.,
2000b; Finneran et al., 2001; Lloyd and Lovley, 2001; Holmes et al., 2002; Jeon et al.,
2004), arsenic (Yamamura et al., 2008), and chromium (Fredrickson et al., 2000a).
132 Electron shuttles have been also employed for biodegradation of nitramine compounds (Kwon and Finneran, 2006, 2008a, 2008b). The cyclic nitramine explosives hexahydro 1,3,5 trinitro 1,3,5 triazine (RDX) and octahydro 1,3,5,7 tetranitro 1,3,5,7 tetrazocine (HMX) are contaminants of concern at many military sites and live fire training installations (Meyers et al., 2007; Mese and Lehmpuhl,
2008; Pennington et al., 2008). Since RDX and HMX are highly oxidized compounds and most contaminated environments are anaerobic, reductive biodegradation of these compounds was investigated in numerous studies (McCormick et al., 1981; Bradley and Dinicola, 2005; Sherburne et al., 2005; Wani and Davis, 2006). In addition, recent studies demonstrated that stimulating Fe(III) reduction via electron shuttling compounds (i.e., anthraqunone 2,6 disulfonate (AQDS) and purified humics) increased RDX and HMX biodegradation at the rate faster than those previously reported (Kwon and Finneran, 2006, 2008b) and further increased the overall extent of mineralization (Kwon and Finneran, 2008a)(also see Chapter 6).
Although electron shuttling mediated RDX reduction is promising, the previous studies (Kwon and Finneran, 2006, 2008b) indicated that a model electron shuttling compound (AQDS) and purified humics reduce RDX and HMX at different rates with slightly different metabolite accumulation dynamics. In addition, AQDS is not cost efficient for the in situ application. Small electron shuttling compounds such as AQDS include a structure similar to aromatic antibiotics, and could be toxic to natural environments (Shyu et al., 2002). For in situ applications, different electron shuttling compounds must be tested to identify an option that is both mechanistically effective and cost efficient.
In this study, several compounds containing quinones, including raw humics extract from mulch and military smoke dyes, were investigated to elucidate whether these compounds can be an alternative source of extracellular electron shuttles for in 133 situ application. The structures of nitramine compounds and several electron shuttling compounds are shown in Figure 8.1. Electron shuttles, such as humic substances, pose no inherent threat to the environment and stimulate contaminant reduction at the extremely low concentrations tested (e.g., 0.05 g/kg purified humics) (Nevin and
Lovley, 2000). Another approach examined military colored smoke dyes derived from quinones and extensively used as signals and markers in both training and combat exercises in the military area (Garrison et al., 1992). Therefore, bioremediation or natural attenuation strategies predicated on catalytic effects of the smoke dyes can be a reasonable approach at military sites.
The objectives of this study are to determine if low cost, easily obtained electron shuttles can be extracted from natural materials, whether these raw humic substances will promote enhanced biodegradation reaction rates, and verify if military smoke dyes can be a viable source of electron shuttles in the military area.
MATERIALS AND METHODS
Mulch extracts. A modified extraction method from the original method
(International Humic Substance Society Method) was used to extract humic substances from mulch. Large fragments of roots and leaves from original material were removed by hand picking and the sample was equilibrated to a pH value between 1 to 2 with 1 M HCl at room temperature. The solution volume was adjusted with 0.1 M HCl to provide a final concentration ratio of 10 mL liquid/1 g dry sample.
The suspension was shaken for 1 hour and then the supernatant was separated from the residue by low speed centrifugation. The mulch residue was neutralized with 1 M
NaOH to pH 7.0 and then 0.1 M NaOH was added under an atmosphere of N 2 to give a final extractant to sample ratio of 10:1. The suspension was extracted under N 2 with intermittent shaking for a minimum of 4 hours. The alkaline suspension was allowed 134 to settle overnight and the supernatant was collected by means of centrifugation. The supernatant was acidified with 6 M HCl with constant stirring to pH 1.0 and then the suspension was allowed to stand for 12 to 16 hours. The material was centrifuged to separate the humic acid precipitate and fulvic acid fractions. The humic acid fraction was re dissolved and neutralized with 1 M NaOH to pH 7.0 and then 0.1 M NaOH was added under an atmosphere of N 2 to give a final extractant to sample ratio of
10:1. The extractant (humic acid fraction) was filtered through a 0.2 m sterilized
PTFE filter into a pre sterilized, anaerobic serum bottle to remove impurities including biomass.
Pure Phase Incubations. Reduced electron shuttling compounds were prepared by sparging the medium bottle with H 2:CO 2 (80:20, vol/vol) in the presence of palladium covered alumina pellets as previously described (Lovley et al., 1999). The reduced shuttles were filtered through a 0.2 m sterilized PTFE filter into a pre sterilized, anaerobic serum bottle to remove cells. The reduced shuttles (200 M of
AH 2QDS, reduced Red9, Red11, and violet1 and 0.2 g/L of reduced purified humics, mulch extract, and henna) were allowed to react with Fe(III) citrate (1mM) or nitrate
(1 mM). Fe(II) concentrations were measured before adding shuttles and in 1, 3, 8, 15
and 30 minutes after adding shuttles. NO 3 reduction was measured before adding shuttles and in 30 minutes after adding shuttles. RDX (60 M) reduction was tested with raw humic extract (0.1 and 0.2 g/L) and Red9 (100 and 200 M).
Experimental tubes (pH 7.8 8.2) were prepared by sparging approximately
4.0ml of 30mM bicarbonate buffer with anaerobic N 2:CO 2 sealing the buffer under an anaerobic headspace. Incubations were performed at 30°C. Samples were collected periodically via anaerobic syringe and needle. To minimize sampling volume, glass inserts (250 L Glass LVI Flat Bottom)(Laboratory Supply Distributors, NJ) were used in the autosampler vials. All experiments were performed in triplicate. 135 Medium and culturing conditions. The basal anaerobic medium consisted of (g l 1 unless specified otherwise): NaHCO 3 (2.5), NH 4Cl (0.25), NaH 2PO 4 H 2O (0.6), KCl
(0.1), modified Wolfe’s vitamin and mineral mixtures (each 10 ml l 1) and 1 mL of
1mM Na 2SeO 4. Fe(III) citrate (45mM) was used as an electron acceptor with the anaerobic medium. The medium was buffered using a 30mM bicarbonate buffer equilibrated with 80:20 (vol/vol) N 2:CO 2, as previously described (Finneran et al.,
2003). Culture tubes/bottles were sterilized by autoclaving at 121 °C under pressure for 20 minutes. Acetate (20 mM) was added as the sole electron donor, depending on the specific culture. Acetate was maintained as sterile, anaerobic, concentrated stock solutions and were added to individual culture vessels using anaerobic, aseptic technique following sterilization. Standard anaerobic and aseptic culturing techniques were used throughout (Lovley and Phillips, 1988). Media were sparged with anaerobic gases passed over a heated, reduced copper column to remove trace oxygen from the gas line. Culture tube/bottle headspaces were flushed with the same gas mixture and sealed under an anaerobic headspace using a thick butyl rubber stopper fastened with an aluminum crimp. All subsequent amendments or transfers were made using sterile needles and syringes that had been flushed with anaerobic gas.
Cultures were maintained in 160 mL anaerobic serum bottles; individual experiments utilized different culture tubes or bottles, as described below.
Resting cell suspension incubations. G. metallireducens was grown anaerobically in freshwater medium with acetate as the sole electron donor and Fe(III) citrate as the sole terminal electron acceptor. One liter of each individual cell culture was harvested during logarithmic growth phase and centrifuged at 5250g for 15 minutes to form a dense cell pellet. Each resultant cell pellet was resuspended in 30 mM bicarbonate buffer under a stream of anaerobic gas. The washed cells were centrifuged again at
136 5250 g for 15 minutes and the resultant biomass was resuspended in 4.0 mL of bicarbonate buffer. Cells were used within 30 minutes of processing.
Experimental tubes were prepared by sparging approximately 5.0 mL of 30 mM bicarbonate buffer with anaerobic N 2:CO 2 and sealing the buffer under an anaerobic headspace. The pH was adjusted between 8.2 and 8.4 because reduced shuttling compounds (hydroquinones) are better reductants at alkaline pH (Uchimiya and Stone, 2006; Kwon and Finneran, 2008a). Raw humic extract (0.04 0.5 g/L) and
Red9 (5 25 M) as electron shuttles were amended from concentrated, anaerobic, sterile stock solutions. An aliquot (0.3 ml) of the resting cells was added to the sealed pressure tubes to initiate each experiment. Samples were collected periodically via anaerobic syringe and needle. To minimize sampling volume, glass inserts (250 L
Glass LVI Flat Bottom)(Laboratory Supply Distributors, NJ) were used in the autosampler vials. All experiments were performed in triplicate.
RESULTS
Fe(III)/NO 3 reduction by reduced electron shuttles
Electron donating capacity of electron shuttling compounds was determined by quantifying the differences in Fe(II) production between with reduced shuttles and with oxidized shuttles. Fe(II) concentration rapidly increased to some extent in 1 minute and remained stable for the remainder of sampling. Reduced AQDS increased the extent of Fe(III) reduction as expected (Figure 8.2A). The Fe(II) concentration accumulated by 200 M of AH 2QDS was nearly stoichiometric. As expected, purified humic acid also showed relatively high reducing capacity. Raw humic extract from mulch, red11 and violet1 also showed good electron donating capacity, which is comparable with purified humic acid. Other reduced shuttles showed some electron donating capacity, but the extents of which were relatively small. 137 AQDS (200 M) reduced 450 M nitrate, which is close to stoichiometric
(Figure 8.2B). Other shuttles also showed very good electron donating capacity to nitrate; in particular, raw humics extract reduced nitrate to the highest extent. As the reduction products, nitrite was not detected (or not detectable because the interference by other peaks), but 10 68 M of ammonium was accumulated.
RDX reduction by reduced raw humics extract and Red9
As the concentration of the extract increased, the rate and extent of RDX reduction increased; approximately 10 M and 20 M of RDX were reduced with 0.1 g/L and 0.2 g/L of reduced mulch extracts, respectively (Figure 8.3A). Oxidized mulch extracts did not reduce RDX. Nitroso metabolites (MNX, DNX, and TNX) were not detected in 20 hours (data not shown). Reduced Red9 (200 M) as well as oxidized Red9 did not reduce RDX in 20 hours (Figure 8.3B).
RDX biodegradation by raw humics extract
Raw humics extract stimulated RDX reduction (Figure 8.4). Increasing the concentration of raw humics extract proportionately increased RDX degradation rates.
The RDX concentration with 0.04 g/l raw humics extract was below the detection limit after approximately 120 hours (data not shown). RDX was completely degraded with 0.1 g/L raw humics extract within 72 hours (Figure 8.4A). RDX with 0.25 or 0.5 g/l of raw humics extract was rapidly degraded in 22 hours. Adding raw humics extract to Fe(III) amended incubations slightly increased the RDX reduction rate
(Figure 8.4B). Very small amount of MNX (< 1 M) accumulated with RDX reduction. DNX and TNX were not detected in 120 hours (data not shown). Raw humics extract in the cell free incubations did not increase RDX reduction rates.
138 RDX and HMX biodegradation by Red9
Red9 increased RDX degradation rates compared to cells alone incubations
(Figure 8.5). RDX degradation rates significantly increased when Red 9 was present; as the concentration of Red9 increases from 5 to 25 M, RDX degradation rates also increased (Figure 8.5A). Adding Red9 to Fe(III) amended incubations slightly increased the RDX reduction rate (Figure 8.5B). Nitroso metabolites (MNX, DNX, and TNX) were not detected in 33 hours (data not shown). Red9 in the cell free incubations did not increase RDX reduction rates.
Red9 also stimulated HMX degradation (Figure 8.6). HMX (ca. 4 M) with 5 or 12.5 M of Red9 was reduced below the detection limit after approximately 33 hours (Figure 8.6A). Adding Red9 to Fe(III) amended incubations significantly increased the HMX reduction rate (Figure 8.6B). Red9 in the cell free incubations did not increase HMX reduction rates. Nitroso metabolite (1 NO HMX) was not detected in 33 hours (data not shown).
DISCUSSION
The results demonstrate that raw humics extract and Red9 can promote nitro group reduction by accepting electrons in microbial respiration and abiotically transferring the electrons to RDX or HMX at rates similar to those tested with purified humics previously (Kwon and Finneran, 2006, 2008b). The electron shuttles are re oxidized and available again for microbial respiration; in this manner the shuttles are catalytic and only a small concentration would be needed to promote these reactions in bioremediation applications.
139 Electron donating capacities by reduced electron shuttling compounds
All compounds containing the quinone functional group showed an electron shuttling capacity to Fe(III) or nitrate, but the extent was different among the compounds (the structures of each electron shuttle investigated in this study are given in Figure 8.1). As expected, reduced AQDS showed the highest electron donating capacity to Fe(III) (Figure 8.2A). The reduced raw humics extract, Red11, and
Violet1 showed relatively high capacity, which is similar to the capacity by reduced humic acid. The Red9 also showed some electron donating capacity, but comparatively small. Some low capacity could be false negatives because the reducing method by palladium with H 2 does not reduce all quinone compounds.
However, the electron donating capacity of each reduced shuttle to nitrate was similar or even better than that of reduced AQDS. These data suggest that all shuttles tested have some potential (even small) as an electron shuttle between cells and Fe(III). In fact, cells (GS15) reduced poorly crystalline Fe(III) hydroxide faster with the smoke dye (i.e. Red9) under growth condition where Red9 act as an electron shuttle, not one way electron donor (data not shown).
Among several potential electron shuttling compounds, raw humics extract and Red9 were further investigated for their capability of RDX and HMX degradation.
Although the electron donating capacity of henna was relatively good, we did not test it further because of its toxicity (Zinkham and Oski, 1996; Raupp et al., 2001; Kök et al., 2004).
Electron donating capacity and rate of reduced shuttles to RDX are significantly lower than to Fe(III). Electron donating capacity of reduced raw humics extract was high for Fe(III) (80 M Fe(III) reduced / 0.1 g/L reduced mulch)(Figure
8.2A), while it was small for RDX (10 M RDX reduced / 0.1 g/L reduced mulch)(Figure 8.3A). Electron donating capacity of reduced Red9 was relatively 140 small, but significant for Fe(III) (30 mM Fe(III) reduced / 200 mM reduced Red9) and
for nitrate (350 mM NO 3 reduced / 200 mM reduced Red9)(Figure 8.2). However,
RDX reduction by reduced Red9 was insignificant (Figure 8.3B).
These data suggest that Fe(III) and nitrate are better electron acceptors than
RDX. These results are consistent with our previous study; the rate of electron
5 transfer from AH 2QDS to Fe(III) is approximately 10 times faster than the rate of
AH 2QDS electron transfer to RDX (Kwon and Finneran 2008c). Therefore, it is possible that reduced raw humics and Red9 also transfer electrons preferentially to
Fe(III) over RDX.
Potential of mulch extract and military smoke dye as electron shuttles for RDX and
HMX degradation
Electron shuttling data generated with model quinones (AQDS and purified humic substances) are critical to understanding electron transfer mechanisms between cells and contaminants. However, for the in situ application the ultimate source of electron shuttle will have to be inexpensive (relative to alternative amendments), readily available, and simple enough that on site personnel, who may not have advanced chemistry backgrounds, can prepare the amendments with little oversight.
Since humic substances are derived from natural organic matter (NOM) it is reasonable to expect that electron shuttle amendments for groundwater amendment will likely be a “raw” form of humic substances.
HS have been extracted from natural materials such as leaf litter, plant and wood compost, or peat material (Nevin and Lovley, 2002a) using a standard method published by the International Humic Substances Society (IHSS) (Swift, 1996). The extractions required to isolate purified humic acids and fulvic acids from natural materials are stringent acid/base extractions with centrifugation and column 141 purification steps (Stevenson, 1982; Swift, 1996). While the extracts are relatively high in quinone content, the extraction procedure is too advanced to be applicable at field sites that may have limited laboratory facilities. One report described a less stringent extraction procedure adapted from the IHSS method that generated usable electron shuttles in very small volumes that increased Fe(III) reduction to varying extents (Nevin and Lovley, 2002a). However, organic and inorganic contaminants were not tested and the “raw” extract was not optimized or scaled up to field systems.
The results presented above were generated with a modified IHSS extraction method in which HF extraction and column purification were omitted. The results suggest that electron shuttles extracted from mulch will stimulate RDX reduction. However, future work will investigate even less stringent extracts; this technology will only gain credibility if electron shuttles that are simply produced in situ are demonstrated to promote biodegradation at rates that are comparable to the purified materials.
Raw humics extract stimulated RDX and Fe(III) reduction in G. metallireducens suspensions (a model Fe(III) and extracellular electron shuttle reducing microorganism) at different rates depending on the concentration of raw humics extract tested (Figure 8.4). Both 0.25 and 0.5 g/l of raw humics extract degraded RDX at the same decay rates (13 nmol RDX h 1 mg cell protein 1)(Figure
8.4A). The difference of degradation rates between these two concentrations was not significant; indicating a threshold concentration between 0.1 g/l and 0.25 g/l under these conditions. Fe(II) was produced more in the presence of raw humics extract suggesting Fe(II) mediated RDX reduction could be possible (Figure 8.4B). The mulch utilized was provided by GSI Consultants (Houston, TX) and was commercially available garden mulch. It was dark brown and brittle, suggesting a high humics organic carbon content. It is likely that other raw humics sources will stimulate RDX reduction at lower concentrations. 142 Red9 showed its capacity to transfer electrons to Fe(III), but it did not transfer electrons to RDX likely because of kinetic limitation (Figure 8.3B). However, resting cell suspension experiments demonstrate that Red9 can act as an electron shuttle to stimulate RDX and HMX reduction (Figure 8.5 and 8.6) suggesting even very low concentration of Red9 can reduce RDX and HMX when it is present as an electron shuttle. Red9 also slightly increased Fe(II) production (Figure 8.5B and 8.6B) suggesting RDX and HMX reduction in the presence of Fe(III) and Red9 can be due to the electron transfer from reduced Red9 directly and from Fe(II) indirectly or both as discussed in the previous study (Kwon and Finneran, 2008c). In spite of limited electron donating capacity of Red9 compared to raw humics extract, the rate of RDX reduction with red9 as an electron shuttle (13.0, 8.4, and 5.6 nmol RDX h 1 mg cell protein 1 with 25, 12.5, and 5 M Red9, respectively) was similar to that with raw humics extract. This result suggest that the small structure of Red9 (Figure 8.1) may be more effective to play a role as electron shuttle than raw humics extract, which has a comparatively larger structure.
Red9 (1 methylamino anthraquinone) is derived from anthraquinone. It has been used in the M18 colored smoke grenade and as ground to ground or ground to air signaling devices, target or landing zone marking devices, and screening devices for unit movements (Garrison et al., 1992). Although its potential toxicity is concern to human health and ecology (Matti Hemmilä, 2007), its electron shuttling role on
RDX and HMX natural attenuation in contaminated areas may be an interest. For example, explosive compounds such as RDX, HMX, and TNT have massively contaminated the Camp Edwards Massachusetts Military Reservation (Clausen et al.,
2004). However, the aquifer material from this area contained low concentrations of these explosives (e.g., less than 1 M; unpublished data). Enhancing natural attenuation by the smoke dye could be possible for this unexpected low level 143 contamination by the explosives since it is also reported that the groundwater in
MMR contained dye compounds, in particular Red9 (Clausen et al., 2004). Therefore, it will be a great interest to investigate whether military smoke dyes can be the source of electron shuttles, and therefore they could stimulate natural attenuation of contaminants in military areas.
144
Figure 8.1. The structures of nitramine compounds and several electron shuttling compounds
145
400 200 M AQDS 0.1g/L Humics 0.1g/L Mulch extract 300 0.1g/L Henna 200 M Red 9
M) 200 M Red 11 200 200 M Violet 1 Fe(II)(
100 A
0 600
500
400 M) ( -
3 300 NO 200 + NH 4 100 B 54 10 68 20 10 25 10 0 Reduced Shuttles Oxidized Shuttles
Figure 8.2. Fe(II) production (A) and NO 3 reduction (B) by several reduced electron shuttles; the numbers in bars indicate the concentration of ammonium ( M) after nitrate reduction. Results are the means of triplicate analyses and bars indicate one standard deviation
146 70
60
50 M)
RDX RDX ( 40
0.2g/L reduced mulch 30 0.1g/L reduced mulch 0.2g/L oxidized mulch A 20
60
50 M)
40 RDX ( 200 M reduced Red9 30 100 M reduced Red9 200 M oxidized Red9 B 20 0 5 10 15 20
time (h) Figure 8.3. RDX reduction by reduced raw humics (A) and reduced Red9 (B). Results are the means of triplicate analyses and bars indicate one standard deviation
147 40
cell+Ac+0.5g/L mulch cell+Ac+0.25g/L mulch 30 cell+Ac+0.1g/L mulch cell+Ac+0.04g/L mulch cell+Ac no cell+Ac+0.04g/L mulch M) 20 RDX ( A 10
0 40 B 1.4
1.2 30 1.0 M) 20 0.8 RDX ( RDX Fe(II)(mM) 0.6 cell+Ac+Fe(III)+0.04g/L mulch 10 cell+Ac+Fe(III) cell+Ac+Fe(III)+0.04g/L mulch (Fe(II)) 0.4 cell+Ac+Fe(III) (Fe(II)) 0 0.2 0 20 40 60 80 time (h) Figure 8.4. RDX reduction by raw humic extracts with (A) or without (B) Fe(III) in the resting cell suspensions of G.metallireducens . Results are the means of triplicate analyses and bars indicate one standard deviation
148 40
cell+Ac+25 M Red9 cell+Ac+12.5 M Red9 30 cell+Ac+5 M Red9 cell+Ac no cell+12.5 M Red9 M) 20 RDX(
10 A 0
cell+Ac+Fe(III)+12.5mM Red9 B 1.4 cell+Ac+Fe(III) 30 cell+Ac+Fe(III)+12.5mM Red9 (Fe(II)) 1.2 cell+Ac+Fe(III) (Fe(II)) M) 1.0 20 RDX( 0.8 Fe(II)(mM) 10 0.6
0 0.4 0 5 10 15 20 25 30 35 time (h) Figure 8.5. RDX reduction by Red9 with (A) or without (B) of Fe(III) in the resting cell suspensions of G. metallireducens. Results are the means of triplicate analyses and bars indicate one standard deviation
149 5
cell+Ac+12.5 M Red9 4 cell+Ac+5 M Red9 cell+Ac no cell+12.5 M Red9 3 M)
HMX ( HMX 2
1 A 0
cell+Ac+Fe(III)+12.5 M Red9 1.2 cell+Ac+Fe(III) 4 cell+Ac+Fe(III)+12.5 M Red9 cell+Ac+Fe(III) 1.0 3 M) 0.8 2 HMX ( HMX Fe(II)(mM) 0.6 1 B 0.4 0 0 5 10 15 20 25 30 35 time (h) Figure 8.6. HMX reduction by Red9 with (A) or without (B) of Fe(III) in the resting cell suspensions of G. metallireducens . Results are the means of triplicate analyses and bars indicate one standard deviation
150 CHAPTER 9
SUMMARY AND CONCLUSIONS
Extracellular electron shuttling compounds increased the rate and extent of
RDX transformation in pure culture incubations and contaminated aquifer material.
Adding extracellular electron shuttling compounds (i.e., AQDS or HS) to the resting cell suspensions ( G. metallireducens ) increased RDX biodegradation and prevented significant accumulation of the nitroso metabolites relative to electron donor alone, which has been the primary strategy to date. A variety of Fe(III) and/or electron shuttle reducing microorganisms (i.e., G. metallireducens, G. sulfurreducens, A. dehalogenans, S. oneidensis , and D. chlororespirans ) reduced RDX or HMX most rapidly with electron shuttle amendments.
The study also investigated extracellular electron shuttle mediated RDX biodegradation and the distribution of ring cleavage metabolites generated by biological degradation (cells) versus the products formed by abiotic degradation
(reduced electron shuttles), and when the two pathways were acting simultaneously.
All pathways were influenced by pH. As pH increased, the rates of RDX reduction by
AH 2QDS also increased. Cells alone reduced RDX faster at the lower pH values.
However, at all pH the rates of the electron shuttle mediated pathways were consistently the fastest, and the proportion of carbon present as formaldehyde, which is a precursor to mineralization, was highest in the presence of electron shuttles.
Approximately 7 – 20 % of RDX was mineralized to CO2 in the presence of cells at
14 all pH tested; AQDS increased the extent of CO 2 produced. Nitrous oxide and nitrite were end products in the strictly abiotic pathway, but nitrite was depleted in the presence of cells to form ammonium. Understanding the different products formed in the abiotic versus biological pathways and the influence of pH is critical to 151 developing mixed biotic abiotic remediation strategies for RDX.
Multiple electron transfer pathways for RDX biodegradation in the presence of bioavailable Fe(III) and electron shuttling compounds were examined using aquifer material incubations, abiotic experiments, and cell suspensions. Electron shuttling compounds stimulated RDX reduction in aquifer sediment. RDX was reduced before onset of significant accumulation of Fe(II) indicating that reduced shuttles transferred electrons to Fe(III) rapidly, and Fe(II) generated reduced RDX. This hypothesis was also supported by kinetics experiments; the rate of electron transfer from AH 2QDS to
5 Fe(III) was approximately 10 times faster than the rate of AH 2QDS electron transfer to RDX. The resting cell suspension ( G. metallireducens ) experiments demonstrated that there are four possible electron transfer pathways for cyclic nitramine biodegradation; however, the rates of the electron shuttle mediated pathways were consistently the fastest. When the Fe(II) mediated electron transfer pathway was inhibited with the Fe(II) ligand Ferrozine, the rate and extent of RDX/HMX degradation decreased, but reduction continued. This suggests that multiple electron transfer pathways overlap in the presence of Fe(III), but inhibiting one pathway does not deter degradation.
Extracellular electron shuttles also stimulate RDX mineralization in RDX contaminated aquifer sediment. RDX loss was similar in electron shuttle amended and donor alone treatments; however, the concentrations of nitroso metabolites, in particular TNX, and ring cleavage products (e.g., HCHO, MEDINA, NDAB, and
+ NH 4 ) were different in shuttle amended incubations. Nitroso metabolites accumulated in the absence of electron shuttles (i.e., acetate alone). Most notably,
14 14 40% 50% of [ C] RDX was mineralized to CO 2 in shuttle amended incubations.
Mineralization in acetate amended or unamended incubations was less than 12% within the same time frame. The other differences in the presence of electron shuttles 152 were the increased presence of NDAB, and the proliferation of Fe(III) reducing microorganisms. NDAB has previously been identified as an aerobic biological metabolite, or product of abiotic RDX transformation. However, we demonstrated that it was also produced biologically by Fe(III) reducing microorganisms, including model pure cultures. RDX was reduced concurrently with Fe(III) reduction rather than nitrate or sulfate reduction. Amplified 16S rDNA restriction analysis (ARDRA) indicated that unique Fe(III) reducing microbial communities ( β and γ proteobacteria ) predominated in shuttle amended incubations. These results demonstrate that indigenous Fe(III) reducing microorganisms in RDX contaminated environments utilize extracellular electron shuttles to enhance RDX mineralization.
A novel species strain MJ1 was isolated from RDX contaminant aquifer material. MJ1 transformed RDX to several metabolites and intermediates under anaerobic condition suggesting that indigenous microorganisms in sediments can degrade RDX effectively. Further characterization of the novel isolate may help RDX bioremediation strategy in situ using the physiological properties of the novel isolate.
For in situ applications, we tested different electron shuttling compounds to identify an option that is both mechanistically feasible and cost efficient. The experiments with the resting cell suspensions of G. metallireducens indicated that
RDX/HMX reduction rate increased with raw humic material extracted from commercial mulch and military smoke dye. The results suggest that these quinone containing compounds could be an alternative source of extracellular electron shuttle for in situ application. Electron shuttles such as raw humics extract from mulch may pose no inherent threat to the environment and stimulate contaminant reduction at the low concentrations. In addition, smoke dyes derived from quinone have been extensively used for military purposes. Therefore, the bioremediation strategies
153 predicated on catalytic effects of these dyes may be a reasonable approach at military sites.
Environmental Engineering Applications
The results demonstrated here will provide unique insight into contaminant transport and aid the development of in situ bioremediation technologies for treatment and long term stewardship strategies addressing subsurface contamination by explosive compounds. However, in situ bioremediation can be effective only where environmental conditions permit microbial growth and activity; its application often depends on the manipulation of environmental parameters to allow microbial growth and degradation to proceed at a faster rate. In addition, the success of in situ biodegradation will be highly site dependant. To control and optimize electron shuttle mediated RDX biodegradation processes, many factors should be considered including the existence of a microbial population capable of degrading the contaminants, sediment type, temperature, pH, electron acceptors, and nutrients. In the following section, electron shuttle mediated biodegradation of RDX is discussed in terms of in situ applications.
The presence of indigenous microorganisms responsible for the contaminant reduction is a key factor in successful bioremediation. This study demonstrates that cyclic nitramine transformation is ubiquitous among the Fe(III) reducing microbial genera most often cited as “environmentally relevant”. Indigenous microbial communities capable of using Fe(III) and electron shuttling compounds is already present at the explosive contaminated sites. Therefore, electron shuttle mediated biodegradation of RDX does not need to introduce exogenous microorganisms into the subsurface.
154 To be successful, electron shuttle mediated bioremediation of RDX requires a procedure for stimulating and maintaining the activity of microorganisms. In situ biodegradation of RDX also requires a continuous supply of a suitable electron donor and shuttle, and a carbon source for energy and cell material. In addition, electron donor and electron shuttle should be added based on the consideration of financial, thermodynamic, and kinetic reasons.
The various kinds of electron donor can be applied in situ . However, there are a number of alternative electron acceptors (e.g., oxygen, nitrate, Fe(III), sufate etc.) that can act as competitors for the electron donor. Therefore, the design accounts for full reduction of these alternative electron acceptors (i.e., concentrations of all electron acceptors that need to be reduced by the electron donor), with the exception of sulfate. Since sulfate reducing conditions may be less favorable for extracellular electron shuttle mediated RDX biodegradation, stoichiometric concentration of the electron donor for not stimulating sulfate reducing bacteria must be added.
The most important factor in electron shuttle mediated RDX biodegradation is to the use of a suitable electron shuttle. The rate and extent of RDX transformed with electron shuttles is much greater than electron donor alone. The study demonstrates that using raw humics extract and military smoke dye creates an efficient, recyclable, cheap, and effective biodegradation system for the decontamination of nitramine compounds. The success of raw humics extract and Red9 in accelerating degradation of RDX suggests that other alternative electron shuttling compounds may also degrade RDX effectively by electron transfer mechanisms mediated by Fe(III) and extracellular electron shuttle reducing microorganisms. There are a variety of options for natural electron shuttles including river sediment humic/fulvic extracts, leaf litter extracts, and possibly other colored dye compounds (e.g., indigo dye) that contain quinones (Rubin and Buchanan, 1985; Nevin and Lovley, 2002a). In the current 155 study, raw humics extract and Red9 increased the rate of RDX reduction relative to cells only controls for G. metallireducens . Similar to purified humics amended incubations (Kwon and Finneran, 2006, 2008b), adding raw humics extract as an electron shuttle stimulated RDX reduction. A slightly lower concentration of raw humics extract, compared to purified HS, was needed for similar RDX degradation.
This suggests that the mulch obtained for this study may have had similar or slightly more humics content to start, and other commercially available sources would likely yield different results. Low molecular weight HS extracted from mulch are soluble and may likely promote the reduction of explosives in situ . In addition, low molecular weight HS are relatively benign in that they are naturally derived and catalytic – i.e. only a low concentration is needed to promote these reactions (Lovley et al., 1998). These data demonstrate the potential for raw HS extracted from mulch to be a source of extracellular electron shuttles, which may be an economic choice for environmental applications.
The success of Red9 in accelerating degradation of both RDX and HMX suggests that other electronegative explosive compounds may be degraded effectively.
Most explosive contaminated sites have a variety of explosive compounds present including TNT, RDX, HMX, and CL 20 (Talley and Sleeper, 1997; Adam et al.,
2004). TNT, a nitroaromatic explosive compound, is reported to degrade easily under reduced conditions in the subsurface (Hawari et al., 2000a); therefore, TNT has rarely been detected in groundwater. The strategy of removing nitro groups in the nitramine explosive compounds by electron transfer via extracellular electron shuttles will be applicable to contaminated sites with other explosive compounds. In fact, our recent experiment indicates that adding electron shuttles to RDX and HMX co contaminated sediments stimulated both RDX and HMX reduction (not published data).
156 The electron donor and electron shuttle can be introduced into a hydraulically controlled system to assure that the microorganisms, donor, and shuttle are in contact with explosive compounds long enough to be utilized. To increase availability of electron donor and shuttle during the reaction, groundwater can be pumped from recovery wells and inoculated with nutrients, electron donor, and electron shuttle, and reintroduced into a upgradient area of the aquifer. Treated groundwater can be returned through injection wells or monitoring wells. The process of a reintroduction procedure will depend on geological settings and geochemical conditions of the sites and local regulations.
The electron donor and electron shuttle can also be introduced as a type of
Permeable Reactive Barrier Wall. For example, a mulch wall can be installed down gradient from the flow path of RDX contaminanted plume. This mulch wall can provide slow distribution of humic material into the contaminant area. In addition, the wall can be mixed with a solid form of electron donor (e.g., chitin as source of acetate). Therefore, the RDX in the contaminated plume may react with the electron donors and shuttles from the barrier to break down RDX into less harmful products.
Cyclic nitramines persist in many subsurface environments despite their perceived “inherently biodegradable” characteristics; more strategies are necessary to promote complete mineralization in situ. Since RDX degrades into a variety of metabolites and intermediates and some products are more toxic than RDX, microorganisms should transform or mineralize RDX to less harmful or harmless products. In addition, understanding the pathways is important for practitioners help refining the technology to better suit low or high bioavailable Fe(III) environments.
The results in this study demonstrate that electron shuttle mediated cyclic nitramine transformation promoted the greatest rate and extent of low molecular weight RDX ring cleavage products, formaldehyde in particular, which indicates that in situ this 157 strategy may lead to more rapid and complete RDX mineralization by targeting electron shuttle reducers and specific biotic abiotic reactions. In fact, sediment incubations indicated that formaldehyde did not accumulate over time and RDX was mineralized greater extent in the presence of electron shuttles. The degradation patterns were consistent when electron shuttles were added.
pH had a marked effect on RDX reduction and metabolite production in the presence of electron shuttle (i.e. AQDS). Since groundwater pH ranges between 6.5 and 8 in general, electron shuttling strategy will work at many sites. However, pH effects on RDX reduction and metabolite production must be considered when designing remediation strategies for the contaminated sites with extreme pH values (<
5); while extracellular electron shuttling ultimately facilitates RDX degradation, pH variation will impact the rate of RDX reduction and metabolite distribution.
In summary, extracellular electron shuttling strategy will provide several advantages for the decontamination of nitramine compounds; quicker cleanup time, low cost, easy of delivery, and risk reduction. The findings in this study can join other bioremediation strategies for RDX in a growing list of available tools for engineered cleanup of explosives contaminated environments. The electron shuttle mediated RDX bioremediation may also help remediation scientists and engineers understand the natural attenuation reactions that contribute to contaminant transformation in anaerobic, subsurface environments.
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175 APPENDIX : CHEMICALS AND ANALYTICAL METHODS
Chemicals . RDX (97% pure) and HMX (99.8% pure) were provided by the U.S.
Army Corps of Engineers, Construction Engineering Research Laboratory (CERL),
Champaign, IL. Hexahydro 1 nitroso 3,5 dinitro 1,3,5 triazine (MNX; 99%), hexahydro 1,3 dinitroso 5 nitro 1,3,5 triazine (DNX; 58% pure with 34% MNX and
8% TNX), hexahydro 1,3,5 trinitroso 1,3,5 triazine (TNX; >99.9%), octahydro 1 nitroso 3,5,7 trinitro 1,3,5,7 tetrazocine ( 1 NO HMX; >98%), methylenedinitramine
(MEDINA), and 4 nitro diazabutanol (NDAB) were purchased from SRI
International (Menlo Park, CA, USA). U [14 C] RDX was purchased from Perkin
Elmer (7.7 mCi/mmol) and dissolved in acetone. Purified humic acid and anthrquinone 2,6 disulfonate (AQDS) were obtained from Sigma Aldrich
(Milwaukee, WI, USA). Disperse Red 9, Disperse Red 11, and Disperse Violet 1 were obtained from Organic Dyestuffs Corporation (East Providence, RI, USA).
Henna was purchased from the online henna tattoo vendor (www.hennaking.com).
HPLC grade methanol was obtained from Aldrich Chemicals. All other chemicals used were of reagent grade quality or higher. Microorganisms . Geobacter metallireducens strain GS 15 (ATCC 53774) and
Geobacter sulfurreducens strain PCA (ATCC 51573) were originally obtained from the University of Massachusetts at Amherst. Anaeromyxobacter dehalogenans strain
K (no ATCC number) and Desulfitobacterium chlororespirans strain Co23 (ATCC
700175) were provided by Dr. Robert Sanford, University of Illinois at Urbana
Champaign. Shewanella oneidensis strain MR1 (ATCC 700550) was obtained from
Dr. Joseph W. Stucki, University of Illinois at Urbana Champaign. Geobacter metallireducens , Anaeromyxobacter dehalogenans , and Desulfitobacterium
176 chlororespirans were maintained using the Ferric Citrate and/or AQDS media, while
Shewanella oneidensis was maintained on aerobic LB medium. Cultures were transferred to fresh medium prior to the onset of any experiments. Analytical techniques . RDX and its metabolites, MNX, DNX and TNX, and HMX and its metabolite, 1 NO HMX, were analyzed using high performance liquid chromatography (HPLC) with a variable wavelength photodiode array (PDA) detector
(HPLC/UV, Dionex) at 254nm as described previously (Fournier et al., 2002). The filtered samples were manually injected or auto injected into a Supelcosil LC CN column (25cm x 4.6mm, 5 m ID) at ambient temperature. A mobile phase consisting of 50% water and 50% methanol was used at a flow rate of 1 mL/min. RDX, MNX,
DNX, TNX, HMX, and 1 NO HMX were compared to certified analytical standards in acetonitrile at known concentrations. MEDINA and NDAB were analyzed using
HPLC at 210nm as described previously (Zhao et al., 2004c). The peaks of MEDINA and NDAB were confirmed by using a reference standard with the HPLC method and retention times were clearly different between two peaks (MEDINA = 6.6 – 6.9 min, NDAB = 7.7 – 7.9 min). In order to double check the production of these compounds, the UV spectrum (200 – 350nm) of these products was compared with that of reference standards. Nitrous oxide was analyzed using gas chromatography
(GC) with thermal conductivity detector (TCD, Hewlett Packard 6890 Series) and
Carboxen 1004 stainless steel micropacked column (1/8 inch x 8 feet; Supelco) that was held isothermally at 100ºC. Helium was used as the carrier gas at a flow rate of
12.5ml/min. The inlet and detector temperature were 195ºC and 250ºC, respectively.
Methanol (CH 3OH) was analyzed using a gas chromatograph (GC) with a flame ionization detector (FID, Hewlett Packard 6890 Series) as described previously
(Monteil Rivera et al., 2005). Nitrite was measured using an ion chromatograph (IC;
177 Dionex 1000) with an AS14A column (250 x 4 mm, Dionex) and isocratic 8mM
Na 2CO 3/1mM NaHCO3 eluent. Ammonium was determined spectrophotometrically at
650nm (Rhine et al., 1998). Formaldehyde was measured by a modified version of
14 14 EPA method 8315A as previously described (Gregory et al., 2004). CO 2 and CH4 were analyzed using gas chromatography (GC; Hewlett Packard 6890 Series) with a gas radiochromatography detector (GC Ram; IN/US system, Tampa, FL) and the
14 14 same column used for nitrous oxide analysis; however, for CO 2 and CH 4 analyses the oven was held isothermally at 120ºC. Helium was used as the carrier gas at a flow rate of 12.8 ml/min. The inlet temperature was 100ºC (Finneran and Lovley, 2001).
Anions in groundwater chemistry were quantified using a Dionex DX 1000 ion chromatograph (Dionex Corp., Sunnyvale, CA) with a Dionex AS4 SC IonPac column. The concentration of reduced AQDS was determined spectrophotometrically at 450 nm versus standards of chemically reduced AH2QDS as previously described
(Lovley et al., 1996b). Aqueous Fe(II) and total solid phase iron concentrations were quantified by the Ferrozine assay as described previously (Lovley and Phillips, 1987).
Each 0.1mL aliquot from the incubations was diluted in 0.5N HCl and 0.1mL aliquot of the diluted solution was added to 4.9mL of 1mM Ferrozine solution, and then quantified at 562nm. Total cellular protein was determined by using the DC Protein
Assay (Bio RAD) and a modified Lowry protein assay (Lowry et al., 1951). pH in aqueous phase was measured by Semi Micro pH probe (Thermo Scientific Inc.).
178 CURRICULUM VITAE EDUCATION University of Illinois at Urbana Champaign ● Doctor of Philosophy in Environmental Engineering and Science – January 2009 Korea University , Seoul, Korea ● Masters of Science in Geochemistry ● Thesis title: Geochemical and Geomagnetic Properties of Drill Core Sediment from Youngjong Island: New Evidences of Sea Level Change during the Late Pleistocene to Holocene (Advisor: Dr. Seong Taek Yun) Korea University , Seoul, Korea ● Bachelor of Science in Earth and Environmental Sciences RESEARCH AND TEACHING EXPERIENCE University of Illinois at Urbana Champaign Graduate Research Assistant (12/2004–present) ● Investigated kinetics and mechanisms of electron shuttle mediate RDX biodegradation with a variety of analytical techniques and materials ● Performed community analysis using molecular genetic tools ● Mentored and trained new students to use analytical instruments and methods University of Illinois at Urbana Champaign Graduate Research Fellow and Research Assistant (8/2003–11/2004) ● Investigated geochemical and microbial variation during well pumping Korea University , Seoul, Korea Graduate Research Assistant (3/1999–2/2000) ● Reconstructed paleoenvironmental change using geochemical and magnetic properties ● Examined occurrence and behavior of fluorine in groundwater Korea University , Seoul, Korea Undergraduate Teaching Assistant (3/1999–12/1999) ● Taught “Geochemistry and Lab” and “Geoenvironment and Experiment” ● Prepared teaching material and lectures ● Graded homework, quizzes, and tests Center for Mineral Resources Research , Seoul, Korea Research Fellow (10/1997–2/2000) ● Pretreated and analyzed environmental samples (partial and total extraction of elements)
179 Korea Basic Science Institute , Seoul, Korea Research Assistant (9/1995 – 8/1997) ● Pretreated environmental samples ● Cleaned and organized laboratory supplies and equipments WORK EXPERIENCE Dong gu Girl’s Middle School , Seoul, Korea Science Teacher (3/2000 – 7/2003) ● Taught basic physics, chemistry, biology, and earth science ● Organized astronomical observation club ● Organized extracurricular activities for students PROFESSIONAL AFFILIATIONS ● American Chemical Society (ACS) ● American Society for Microbiology (ASM) ● Geological Society of America (GSA) SCHOLARSHIPS AND HONORS ● Student Research Contest (2 nd place), Geosyntec consultants (3/2008) ● Racheff Travel Award, Civil and Environmental Engineering, University of Illinois (9/2006, 3/2005) ● Corporate Activities Program Student Travel Grant, American Society for Microbiology (5/2006) ● Graduate School Student Travel Award, University of Illinois (10/2005, 10/2004) ● Student Travel Award, American Society for Microbiology, (8/2005) ● University Fellowship, University of Illinois (8/2003 – 7/2004) ● Han Shin Scholarship Award, Han Shin Foundation, (8/1996 – 2/1999) ● Brain Korea (BK) 21 Scholarship Award, Ministry of Education (3/1999 – 2/2000) ● Korea Exchange Bank Scholarship Award, Korea Exchange Bank (3/1998) ● Honor Scholarship Award, Korea University (3/1996) PEER REVIEWED PUBLICATIONS ● Kwon, M.J. , S.T. Yun, S.J. Doh, B.K. Son, and K.S. Choi, 2008. Metal Enrichment and Magnetic Properties of Core Sediment from the Eastern Yellow Sea: New Insights of Sea Level Change During the Glacial/Holocene Transition. Quat. Int. in revision
180 ● Kwon, M.J. , and K.T. Finneran, 2008. Hexahydro 1,3,5 trinitro 1,3,5 triazine (RDX) Reduction is concurrently catalyzed by Direct Electron Transfer from Hydroquinones and from the Hydroquinone Fe(III)/Fe(II) couple during Electron Shuttle Amended Biodegradation. Environ. Eng. Sci. accepted ● Zhang, Y., K. Reinauer, K. Dunett, M.J. Kwon , and K.T. Finneran, 2008, Hydrogenophaga carboriundus sp. nov., a Tert butyl Alcohol Oxidizing, Psychrotolerant Aerobe Isolated from Granular Activated Carbon (GAC), Int. J. Syst. Evol. Microbiol. , in revision . ● Kwon, M.J. , and K.T. Finneran, 2008. Biotransformation Products and Mineralization Potential for Hexahydro 1,3,5 trinitro 1,3,5 triazine (RDX) in Abiotic versus Biological Degradation Pathways with Anthraquinone 2,6 Disulfonate (AQDS) and Geobacter metallireducens , Biodegradation , 19(5), 705 715. ● Kwon, M.J. , R.A. Sanford, J. Park, M.F. Kirk, and C.M. Bethke, 2008. Microbiological Response to Well Pumping, Ground Water, 46(2), 286 294. ● Kwon, M.J. , and K.T. Finneran, 2008. Hexahydro 1,3,5 trinitro 1,3,5 triazine (RDX) and Octahydro 1,3,5,7 tetranitro 1,3,5,7 tetrazocine (HMX) Biodegradation Kinetics Amongst Several Fe(III) Reducing Genera, Soil Sediment Contam. , 1 7(2) ,189 203. ● Kwon, M.J. , and K.T. Finneran, 2006. Microbially Mediated Biodegradation of Hexahydro 1,3,5 trinitro 1,3,5 triazine by Extracellular Electron Shuttling Compounds, Appl. Environ. Microbiol., 72(9) , 5933 5941. ● Chae, G.T., S.T. Yun, M.J. Kwon , Y.S. Kim, and B. Mayer, 2006. Batch Dissolution of Granite and Biotite in Water: Implication for fluorine geochemistry in groundwater, Geochemical Journal , 40 , 95 102. MANUSCRIPTS SUBMITTED AND IN PREPARATION FOR PUBLICATION ● Kwon, M.J. , and K.T. Finneran, 2008. Hexahydro 1,3,5 trinitro 1,3,5 triazine (RDX) Mineralization and Biological 4 nitro 2,4 diazabutanal (NDAB) Production in the presence of Extracellular Electron Shuttling Compounds and a Unique Fe(III) Reducing Microbial Community, submitted to Environ. Sci. Technol. ● Kwon, M.J. , and K.T. Finneran, 2008. Characterization of a Hexahydro 1,3,5 trinitro 1,3,5 triazine (RDX) degrading, novel anaerobe isolated from contaminated aquifer material. ● Kwon, M.J. , B. Powers, and K.T. Finneran, 2008. Mulch Extract and Military Smoke Dye as Extracellular Electron Shuttles for Biodegradation of Hexahydro 1,3,5 trinitro 1,3,5 triazine (RDX).
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