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a Environmental Sciences Division, Oak Ridge National Laboratory, Oak Ridge, TN 37831
b Faculty of Earth System and Environmental Sciences, Chonnam National University, 300 Yongbong-Dong, Buk-Gu, Gwangju, 500-757, Korea
c Environmental Engineering Research Centre, School of Architecture, Civil and Environmental Engineering, and Planning, Queen's University of Belfast, Belfast BT9 5AG, Northern Ireland, UK
* Corresponding author (rohy{at}chonnam.ac.kr)
Received for publication April 11, 2005.
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ABSTRACT |
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 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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Abbreviations: BCV, Bear Creek Valley • CBD, citrate–bicarbonate–dithionate • DOE, Department of Energy • DXRD, differential X-ray diffraction • ECEC, effective cation exchange capacity • EDX, energy dispersive X-ray • FRC, field research center • HIV, hydroxy-interstratified vermiculite • IC, ion chromatograph • MRB, metal-reducing bacteria • NABIR, Natural and Accelerated Bioremediation Research • ORR, Oak Ridge Reservation • PCB, polychlorinated biphenyl • PCE, tetrachloroethylene • RIR, reference intensity ratio • SEM, scanning electron microscope • TIC, total inorganic carbon • TOC, total organic carbon • XRD, X-ray diffraction • XRF, X-ray fluorescence
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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The site is located in Bear Creek Valley (BCV) of the ORR and is on a subcrop of the Nolichucky Shale, a member of the Middle and Late Cambrian Canansauga Group formation. Past waste disposal activities at the pond site have created a mixed-waste plume of contamination including organic site-related contaminants such as endrin aldehyde, methoxychlor, polychlorinated biphenyl (PCB), 2-hexanone, 4-methyl-2-pentanone, methylene chloride, tetrachloroethylene (PCE), and toluene in the underlying unconsolidated residuum and competent shale bedrock (Watson et al., 2001; www.esd.ornl.gov/nabirfrc/; verified 5 Sept. 2005). The plume extends >150 m deep beneath the ponds and extends about 1 km along geologic strike both east and west of the ponds. Remedial activities of the site have been ongoing since 1984, when pond wastes were partially neutralized using carbonate sediment and biologically denitrified using acetate. The site was capped and paved thereafter (Cook et al., 1996). Where dense liquid wastes were disposed, contamination of deep ground water and sediments in the Nolichucky Shale has also occurred.
If left untreated, these contaminated subsurface media may represent human health hazards or serve as a potential source of contamination for underlying aquifers during natural leaching episodes (Elless and Lee, 2002). Recent research by Natural and Accelerated Bioremediation Research (NABIR) investigators has shown that there are several enhancing or inhibiting factors of bacterial U(VI) reduction by metal-reducing bacteria (MRB). The microbial reduction of U(VI) precipitates solid U(IV) phases as microbial reduction of uranyl carbonate and precipitation of uraninite (UO2) has the potential to restrict the risk and the transport of uranyl-carbonate in aqueous environments by producing solid sparingly soluble phases. Microbial U(VI) reduction is inhibited by the presence of competitive electron acceptors such as nitrate (Senko et al., 2002; Finneran et al., 2002a, 2002b; Elias et al., 2003; Istok et al., 2004) and Fe(III) oxyhydroxides (Wielinga et al., 2000) as well as by the presence of geochemical oxidants such as manganese oxides (Liu et al., 2002; Fredrickson et al., 2002). The presence of Ca inhibits microbial reduction of U(VI) at neutral pH due to the formation of uranyl carbonate complexes, Ca2UO2(CO3)3 (Brooks et al., 2003; Gu et al., 2003). Other factors including low pH and high concentrations of toxic metals and/or co-contaminants such as Al and Ni may inhibit microbial reduction of U(VI) in the FRC sites (Gu et al., 2003). The microbial reduction of U(VI) is enhanced by addition of a C source and nutrients (Anderson et al., 2003; Finneran et al., 2002a; Holmes et al., 2002; Istok et al., 2004; Petrie et al., 2003; Suzuki et al., 2003; Shelobolina et al., 2003) and natural organic matter (Gu and Chen, 2003). The NABIR investigators have also elucidated native metal reducing bacterial communities in the subsurface media contaminated with radionuclides and heavy metals (Anderson et al., 2003; Chang et al., 2001; Holmes et al., 2002; Petrie et al., 2003).
Physicochemical and mineralogical characterization studies of the contaminated media are rarely examined on contaminated geological material because of the cost and time expenditure. However, without this prior knowledge of the soil mineralogy and media characteristics, it may be hard to select remediation strategies for contaminated soils, particularly those contaminated with U, as the solubility of U is mediated by redox reactions and by its association with other elements (Roh et al., 2000). By accurately assessing the contaminant nature rather than the extent of the contamination, the development and remediation technology may proceed by logical approaches (Elless and Lee, 2002).
The research at the FRC has been a part of a larger study involving the in situ immobilization of U(VI) and Tc(VII) by microbial reduction to sparingly soluble U(IV) and Tc(IV) species. Preliminary analyses indicated that factors other than electron and nutrient availability were limiting or inhibiting microbial growth at the site, and these factors include low pH and high concentrations of co-contaminants such as Al, Ni, and nitrate (Gu et al., 2003). Bioremediation using microbial reduction of metals may be feasible if mineralogical bioavailability and toxicity is characterized. The objective of this study was to provide improved characteristics of the nature of contaminants and co-contaminants in the subsurface media of soils and deeply weathered saprolite to assess the feasibility of bioremediation.
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MATERIALS AND METHODS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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An undisturbed core (3.8 cm diam.) was collected to a depth of about 9 m with a pneumatic hammer-driven coring device (Fig. 1). Polyurethane core tubes were cut to 15.2 cm and sealed with plastic caps immediately after removal. Characteristics (i.e., color) of the soil–saprolite core material were described according to Soil Survey Staff (1992). The soil–saprolite samples were air-dried and crushed, and then screened through a 2-mm sieve. The sieved samples were used for physicochemical and mineralogical characterization.
Chemical Characterizations of Soil–Saprolite
pH and Effective Cation Exchange Capacity
Soil–saprolite pH was measured in 1:1 soil/water (w/w basis) slurry and in 1:1 soil/10 mM CaCl2 solution to approximate ionic strength of normal soil solution (McLean, 1982) using a 920A pH meter with 8103 Ross probe (ThermoOrion, Boston, MA). Effective cation exchange capacity (ECEC) was determined using 0.2 M NH4Cl solution and 1 g soil–saprolite (Sumner and Miller, 1996). The supernatant was measured by ICP-MS (Thermo Jarrell Ash, Inductively Coupled Plasma–PolyScan Iris Spectrometer, Franklin, MA).
Oxalate, Citrate–Bicarbonate–Dithionate, and Nitric Acid Extraction
Oxalate extraction dissolved amorphous iron oxides, which could contribute significantly to the physical and chemical properties of soils, since they may have high CEC, high surface area, and high reactivity. It was desirable to characterize and quantify the amount of amorphous iron oxides as well as that of the crystalline component. To reveal the amount of amorphous iron oxides, the sieved soil–saprolite (<2 mm, 250 mg) was extracted with 50 mL of a 0.2 M ammonium oxalate solution adjusted to pH 3.0 with concentrated HCl for 2 h in the dark and washed twice with 25 mL of 0.5 M ammonium carbonate and 50 mL deionized water (Jackson et al., 1986). The supernatants were analyzed using the same ICP-MS procedure.
For the CBD extraction of crystalline iron oxides, 1 or 0.5 g (according to the clay content) of the sieved soil–saprolite was mixed with 40 mL of 0.3 M sodium citrate solution and 5 mL of 1 M sodium bicarbonate and 2.0 g sodium dithionate (Jackson et al., 1986) at 80°C for 30 min with vigorous stirring. After twice washing with 40 mL of sodium citrate solution, supernatants were analyzed by the same method as that used for the oxalate extraction.
Uranium was extracted by shaking 0.5 g of core material in 15 mL of 2 M HNO3 (three sequential extractions) overnight and then centrifuging at 3500 rpm for 30 min. The supernatant was mixed with 250-mL aliquot of deionized water and analyzed as above.
Mineralogical Characterization of Soil–Saprolite
Mineral assemblages of bulk samples were identified by X-ray diffraction (XRD) analysis with Cu-K
radiation (40 kV/30 mA) using divergent and scattering slits of 1 mm, a receiving slit of 0.15 mm, 0.02° 2
steps, and a counting time of 2 s per step. The differential X-ray diffraction (DXRD) patterns of the bulk samples were obtained after oxalate- and CBD-extraction (Jackson et al., 1986). For the semiquantification of major mineral constituents, the reference intensity ratio (RIR) method was used (Chung, 1974), in which the equation between peak intensity and the constituent fraction of each mineral could be obtained. The representative samples were compared with the results of chemical compositions of bulk samples using X-ray fluorescence (XRF).
Selected samples of soil–saprolite were dried at room temperature and characterized by scanning electron microscopy (XL30FEG SEM, Philips, Eindhoven, the Netherlands) equipped with an energy dispersive X-ray (EDX) analyzer.
Chemical Characterization of Ground Water
Ground water samples were collected from screened wells and filtered using 0.45-µm in-line filters. Ground water pH and Eh, without exposing to the air, were measured in line with a precalibrated multiparameter probe (XL6000M, Yellow Springs Instruments, CO). Unacidified samples were used for the analysis of major anions in the ground water by means of an ion chromatograph (IC) equipped with a conductivity detector (DX-120, Dionex, Sunnyvale, CA). Alternatively, sulfate was determined with the methylene blue method and spectrophotometer (DR/2000, Hach, Loveland, CO) while nitrate and nitrite were analyzed using colorimetric test kits. Aliquots of the unacidified samples were used for the analysis of total organic carbon (TOC) and total inorganic carbon (TIC) using a total organic carbon analyzer (TOC-5000A, Shimadzu, Tokyo). The TIC was then converted to bicarbonate or carbonate concentrations in ground water. Alkalinity was determined by titration according to Standard Methods (APHA). Acidified ground water samples were used for elemental analysis such as Ca, Mg, Fe, Al, Mn, Ni, Na, K, U, and transition metals using the same ICP-MS procedure.
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RESULTS AND DISCUSSION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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The soil–saprolite pH varied from 4.6 to 10.7 (1:1 H2O) (Fig. 2 ). The pH values of the soil–saprolite cores measured with 10 mM CaCl2 solution were slightly lower (4.2–10.7) than pH measured with H2O because of a higher displacement of H+ and Al3+ ions. The pH of the fill material (up to 6 m BLS) ranged from 7.0 to 10.5, whereas the pH of the underlying saprolite (6–8 m BLS) ranged from 4.5 to 8. The higher pH values of the fill material were attributed to their consisting of calcareous limestone at depths between 390 and 516 cm BLS with pH > 8.5 suggesting that the higher pH might be due to the neutralization and leaching of limes (Kang et al., 2001).
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The trend of ECEC variation was similar to that of pH. Values of ECECw/o Ca and ECECother (data not shown) showed the opposite trend of the ECEC variation;ECECw/o Ca and ECECother showed high values from 38.6 to 83.2 and 32.5 to 81.0 cmolc kg–1 in the low pH sample (623–672 cm BLS) as well as from 3.8 to 5.6 and 4.2 to 4.8 cmolc kg–1 at 766 to 802 cm BLS.
Oxalate, Citrate–Bicarbonate–Dithionate, and Nitric Acid–Extractable Elements
Radionuclides
Results obtained by oxalate, CBD, and nitric acid extraction showed considerable variation in the concentration of radionuclides and toxic metals depending on the reagents used (Table 1). Low concentrations of Sro were detected in amorphous iron oxide throughout the core samples (data not shown), supporting the view that Sr could substitute for Ca in carbonate minerals as a result of their same charge and similar ionic radius (Blundy and Wood, 1991) and would have been dissolved by the low pH contaminated water in the subsurface environments.
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Metals
The S-3 Ponds site served as a source of contaminating Cu, Pb, Ni, and Zn in the soils and Cd, Cr, Co, Cu, Pb, Ni, and Zn in the ground water (Watson et al., 2001). This study was focused on five toxic metals (Cr, Co, Ni, Zn, and Pb) and unselected Cu and Cd due to their variation of data. Apparent concentrations of selected metals in soil–saprolite varied according to the reagents used (Table 1). Although there was difference among the elements, concentrations of metalc were higher than those of metalo because citrate should remove all oxalate-extractable Fe in the soil–saprolite from a NABIR Field Research Site. Figure 3b shows HNO3 (2 M) was not the best extraction method to extract metals coexisting with iron oxides.
High concentrations of Cr, Cu, and Pb were detected at shallower depths than 390 cm and decreased with further depth reflecting the contamination event at the surface. When metals have been derived from anthropogenic sources, these extractable metals exhibited a higher correlation with extractable Fe than total Fe including the Fe in crystal structures (Moon et al., 2001).
Various correlation coefficients were shown as follows: Feo–Cro (r2 = 0.73), Feo–Nio (r2 = –0.34, negative means inverse proportion), Feo–Coo (r2 = –0.07), Feo–Zno (r2 = 0.73), Feo–Pbo (r2 = 0.35), Mno–Nio (r2 = 0.49), Mno–Coo (r2 = 0.80) for oxalate; Fec–Crc (r2 = 0.11), Fec–Nic (r2 = 0.13), Fec–Coc (r2 = 0.33), Fec–Znc (r2 = 0.56), Fec–Pbc (r2 = 0.10), Mnc–Nic (r2 = 0.27), Mnc–Coc (r2 = 0.78) for CBD extraction. Two dominant trends emerged: (i) Cr, Zn, and Pb correlated with amorphous iron oxide phase rather than crystalline phase, (ii) Co and Ni were proportional to Fec and Mno, (iii) correlations between Fe (or Mn) and other metals in the amorphous phase were much higher than in the crystalline phase implying anthropogenic metal contamination.
Major Elements
Major elements in the soil–saprolite also showed considerable variation of concentration depending on the reagents used (Table 1). The largest Mno and Uo peak concurrently appeared at 633 and 613 cm BLS and the second peak of Alo (766–802 cm), which showed bimodal shape, coincided with the second peak of Uo in the range of 752 to 802 cm (Fig. 3c). On the other hand the concentration levels of Feo decreased with increased depth. Manganesec and Mnn showed similar trends as Uc and Un, respectively. Ironc and Alc did not show trends when compared with Uc. However, Cac was inversely related to Uc. This trend also occurred between Can and Un. This result complies well with extractable U decreasing significantly in the presence of free carbonates (Brooks et al., 2003).
The ratio of Feo/Fec-o dramatically decreased at about 600 cm BLS followed by a depletion of amorphous Fe below this "transition zone" (Watson et al., 2001), which exhibited cracks and seams and rapid preferential flow (Jardine et al., 2003). This depletion of amorphous Fe coincides with low pH and high U concentration in this zone (Fig. 3d).
Correlation of major elements in each extraction is as follows: Alo–Feo (r2 = 0.16), Alo–Uo (r2 = 0.13) and Feo–Cao (r2 = 0.41); Alc–Cac (r2 = 0.43), Alc–Fec (r2 = 0.37), Alc–Uc (r2 = 0.39) and Fec–Mnc (r2 = 0.15). On the contrary, Alo–Mno (r2 < 0.01), Feo–Mno (r2 < 0.01), Feo–Uo (r2 < –0.01), Alc–Mnc (r2 = 0.08), and Fec–Uc (r2 = 0.04) relations were poor. If U was incorporated into iron hydroxide, it should show mobility like metals and have good relation with Feo. The dominant contaminant U and Mn showed a low correlation with amorphous Fe such as Feo–Mno (r2 = 0.01) and Feo–Uo (r2 < –0.01). This suggests that the existence and mobility of U and Mn are different from those of Fe and other metals and this may affect the mobility of Mno and Uo.
Calcium inhibits the bacterial U(VI) reduction of U(VI) to U(IV) (Brooks et al., 2003). The inhibitive role of Ca is supported by the fact that Uc–Cac had an evident negative relation (r2 = –0.36) with r2 < 0.01 of Uo–Cao. This reflects that at least U precipitated or became associated to crystalline iron oxide phase was blocked by Ca as well as Ca precipitate instead of U.
The averaged ratio of the concentrations of oxalate-extraction over that of CBD-extraction with maximum and minimum values showed that Cr, Co, Cd, Pb, and U appeared to coexist with the amorphous Fe phase rather than the crystalline iron oxide phases (Fig. 4 ). Because CBD treatment extracts not only crystalline iron oxide but also amorphous iron oxide, which can be dissolved in oxalate, in turn, the same dissolved amounts derived from these two reagents indicate that those elements were related to the Fe amorphous phase. It is another advantage for metal-reducing bacteria to use the amorphous Fe phase as an electron acceptor because Fe(III) reducing bacteria may sequester inorganic contaminants including Cr and Co into more stable forms (Roh et al., 2001).
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The whole ground water data of Area 2 showed that the maximum concentrations of U (<8.0 µM) and NO3– (<124 mM) were lower than concentrations of them at Area 1 (U < 32 µM; NO3– < 199 mM) and Area 3 (U < 256 µM; NO3– < 803 mM) (www.esd.ornl.gov/nabirfrc/; verified 5 Sept. 2005). Ground water analyses indicated that factors other than electron donor and nutrient availability are favoring microbial growth at Area 2, and these factors likely include near neutral pH (about 6) and low concentration of a co-contaminant, nitrate (Table 2).
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Based on the thermodynamic consideration, Fe(OH)3(a) (0.68), barite (0.65), and celestite (1.96) could be controlling factors in ground water of FW200 borehole by saturation index; in contrast, manganite (–5.80) and uraninite (–3.11) have low potential. The Fe(OH)3(a) is a conclusively controlling factor and Mn species did not participate in it. Adsorption and substitution of metals into Fe(OH)3(a) can explain the mobility of toxic metals (Cr, Zn, and Pb), which showed good correlation in Feo fraction.
Based on the correlation coefficients, amorphous Fe can retain toxic elements and Al–U showed a good relationship. Although poor relationships exist between Fe–U, Fe–Mn, Mn–U (Mno–Uo r2 = 0.03, Mnc–Uc r2 = 0.03), Mn and U, which are under the inhibition of Ca, apparently show similar trends of mobility and resulted in the possibility of different mechanisms of Mn and U mobilities beyond coprecipitation or incorporation with Fe phases.
The similar redox intensity was found to be one reason for the trends of U and Mn mobility. That of Fe is far from Mn and U. The sequence of reduction reactions in the order of decreasing redox intensity at pH 7 is as follows (Stumm and Morgan, 1981):
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Based on recent monitoring (4 Mar. 2004) a field test of 18 monitoring wells showed Eh values ranged from –129 to 375 mV (peo, –2.20 to 6.37) and averaged 170 mV (pe = 2.89). This revealed that nitrate, U, and Mn but not Fe have the potential to take part in geochemical reduction reactions (Eq. [1]
–[3]).
Considering the Mn ratio near unity (Fig. 4), a considerable amount of Mn should have incorporated or co-precipitated with amorphous iron oxides; however, Feo–Mno (r2 = 0.01) showed little correlation. When Mn2+ was released to aqueous solution during weathering, it should be more stable toward oxidation than is Fe2+ (Hem, 1992). If Mn is in contact with the atmosphere, it can precipitate as a crust of Mn4+ oxide; however, the pH–Eh condition of this site (Fig. 5 ) showed that Mn2+ is stable in the ground water. This can be a factor in the form of an amorphous manganese oxide phase from Mn2+ just above the interface zones of reclaimed soil–saprolite and saprolite–bedrock.
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Further studies that stimulate microbial U(VI) reduction in high nitrate-contaminated sites may be in order. The nitrate concentration of Area 2 is another advantage of applying remediation to Area 2.
X-Ray Diffraction Analysis and Scanning Electron Microscope Observation
The mineral assemblages of soil–saprolite samples are similar, but they can be classified into two groups by the relative contents of each mineral representing reclaimed soil and weathered saprolite (Fig. 6
). The reclaimed soils (up to 6 m BLS) had larger amounts of quartz, anorthite, calcite, and dolomite (81.8% vs. 44.5% of saprolite as a nonnormalized average) whereas saprolites (6–8 m BLS) had larger amounts of microcline and clay minerals such as illite, kaolinite, and vermiculite (51.1% vs. 18.3% of reclaimed soil). Differential XRD patterns also showed different extractable mineral phases. Calcite and dolomite were dominantly extracted from reclaimed soil zone (at 418 cm BLS) and goethite, ferrihydrite, hydroxy-interstratified vermiculite (HIV), and vermiculite from saprolite zone (at 778 BLS). The most abundant clay mineral, illite, showed variation with depth and the dominant peaks of the relative content occurred in each middle part of reclaimed soil and saprolite zones (Fig. 7
), and small ones at about 600 and 800 cm coincided with the peaks of low pH, high ECECw/o Ca, and Uo. This indicates that the mineralogy of a site can be used to predict the distribution or retardation of contaminants. Historic continuous acidic weathering dissolved even clay minerals as weathering products derived from feldspar. Based on this geochemical study and others (Gu et al., 2003), the high correlation between extractable U–Al was due to the Al-oxyhydroxide weathered from clay minerals and feldspars and it was reflected in the small relative amount of illite. This chemical and mineralogical site characterization study agreed with the thermodynamic result that sorption and coprecipitation of U and Tc with amorphous Al and Fe oxyhydroxides could be a major mechanism responsible of their removal at pH below about 5.5 (Gu et al., 2003). They found that basaluminite [Al4(OH)10SO4] controls activity of Al3+ below pH 7.7, whereas Al(OH)3(a) controls the activity of Al3+ at higher pH. However, the soil at around the 650- and 800-cm depth, which have high concentrations of U, showed low pH; therefore, high correlation between U and Al might be controlled by basaluminite weathered from artificial acidic weathering of feldspars and clay minerals (Fig. 8
).
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To examine the quality of the data using XRD, we correlated the chemical compositions of bulk samples using XRF with the mineral composition calculated from XRD by using the ideal composition shown on the PDF file of each mineral. Though the contents such as Fe2O3 and MnO, which were calculated by XRD analysis, showed lower values than those calculated by chemical analysis, the r2 value was 0.96. The difference was likely due to minerals like ferrihydrite, goethite, and Mno phase in the bulk sample.
The lack of Sro is an evidence of intensive leaching of anorthite, because Sr usually substitutes for Ca in feldspar (Blundy and Wood, 1991; Ren, 2004). Calcium including Sr leached from feldspar and Sr was not detected in the amorphous phase, indicating the stabilization into crystalline phase like celestite (SrSO4) based on thermodynamic consideration. SEM and EDX analysis of samples obtained at 662 cm BLS showed anorthite (Fig. 8a) in combination with clay minerals such as illite and vermiculite as filling or coating materials (Fig. 8b).
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CONCLUSIONS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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Examined site-specific characteristics were (i) large fractions of radionuclides (U and Ba) and metals (Cr, Ni, and Pb) in the amorphous phase, which may be easily used by metal-reducing bacteria during soil development of reclaimed soil, in turn, that condition indirectly leads to low toxic environment for microbes; (ii) the similar trend of mobilities of Mn and U, indicating that their speciation is dominantly affected by not only redox potential, but also abiotic geochemical adsorption or coprecipitation; (iii) site specific mineralogical aspects such as low relative content of clay minerals and high ECEC and Fe or Al oxyhydroxide in the U-rich zone, which can cause adsorption of U(VI); and (iv) low radionuclide contamination levels (= low toxicity), low nitrate level (= low interference of easily reducible phase), and more neutral ground water pH compared with Areas 1 and 3, FRC sites.
This milieu of subsurface furnishes good conditions for microbial growth and with appropriate addition of electron donors and nutrients bioremediation of U by metal-reducing microorganisms may be a feasible remedial alternative. Bioremediation may also enhance contaminant immobilization through the decreased solubility, and leachability by transforming amorphous phases to crystalline oxide phases.
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ACKNOWLEDGMENTS |
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REFERENCES |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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