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Anaerobic bioremediation of hexavalent uranium in groundwater by reductive precipitation with methanogenic granular sludge
water research 44 (2010) 2153–2162
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Anaerobic bioremediation of hexavalent uranium in groundwater by reductive precipitation with methanogenic granular sludge Aida Tapia-Rodriguez, Antonia Luna-Velasco, Jim A. Field*, Reyes Sierra-Alvarez Department of Chemical and Environmental Engineering, University of Arizona, P. O. Box 210011, Tucson, AZ 85721, USA
article info
abstract
Article history:
Uranium has been responsible for extensive contamination of groundwater due to releases
Received 31 August 2009
from mill tailings and other uranium processing waste. Past evidence has confirmed that
Received in revised form
certain bacteria can enzymatically reduce soluble hexavalent uranium (U(VI)) to insoluble
16 December 2009
tetravalent uranium (U(IV)) under anaerobic conditions in the presence of appropriate
Accepted 17 December 2009
electron donors. This paper focuses on the evaluation of anaerobic granular sludge as
Available online 28 December 2009
a source of inoculum for the bioremediation of uranium in water. Batch experiments were performed with several methanogenic anaerobic granular sludge samples and different
Keywords:
electron donors. Abiotic controls consisting of heat-killed inoculum and non-inoculated
UASB
treatments confirmed the biological removal process. In this study, unadapted anaerobic
Uranium reduction
granular sludge immediately reduced U(VI), suggesting an intrinsic capacity of the sludge
Reductive precipitation
to support this process. The high biodiversity of anaerobic granular sludge most likely
Radionuclide
accounts for the presence of specific microorganisms capable of reducing U(VI). Oxidation
Bioreduction
by O2 was shown to resolubilize the uranium. This observation combined with X-ray
Uraninite
diffraction evidence of uraninite confirmed that the removal during anaerobic treatment was due to reductive precipitation. The anaerobic removal activity could be sustained after several respikes of U(VI). The U(VI) removal was feasible without addition of electron donors, indicating that the decay of endogenous biomass substrates was contributing electron equivalents to the process. Addition of electron donors, such as H2 stimulated the removal of U(VI) to varying degrees. The stimulation was greater in sludge samples with lower endogenous substrate levels. The present work reveals the potential application of anaerobic granular sludge for continuous bioremediation schemes to treat uraniumcontaminated water. ª 2009 Elsevier Ltd. All rights reserved.
1.
Introduction
Interest for uranium remediation has increased due to the growing awareness of contamination at mining and processing sites. The environmental contamination results from the
perturbation of naturally occurring uranium minerals through mining and processing in the nuclear fuel cycle as well as during phosphate enrichment. Mill tailings from mining are one of the major causes of extensive uranium contamination due to the large volume of tailings and previous lack of
Abbreviation: VSS, Volatile suspended solids. * Corresponding author. Tel.: þ1 (520) 626 5858; fax: þ1 (520) 621 6048. E-mail address: [email protected] (J.A. Field). 0043-1354/$ – see front matter ª 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2009.12.030
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regulations for their disposal (USEPA, 1995). Uranium is sometimes accompanied by nitrate and sulfate contamination due to the use of nitric and sulfuric acids to extract uranium (Abdelouas, 2006), as well as by the oxidation of primary sulfide minerals from acid mine drainage (USEPA, 1995). The primary public health concern of uranium is its chemical toxicity leading to kidney diseases (ATSDR, 1999). The U.S. Environmental Protection Agency (EPA) has set the primary drinking water limit for uranium at 30 mg L1, which protects the public from kidney toxicity according to the results obtained in epidemiological studies (WHO, 2004). Hexavalent uranium (U(VI)) and tetravalent uranium (U(IV)) are the most common valence states of uranium in nature (Fredrickson et al., 2000). U(VI) predominantly occurs in the uranyl ion form (UO2þ 2 ) (Langmuir, 1978), which is the most reactive and mobile form due to its solubility in water and tendency to form complexes with ligands present in natural waters such as carbonates (Abdelouas, 2006). U(IV) is a highly stable and insoluble form of uranium that generally occurs as the mineral uraninite (UO2(s)). The levels of uranium in the environment are, strongly dependent on pH and redox properties of the subsurface environment. High removal efficiencies (>90%) can be attained by conventional treatment of uranium contamination based on physical-chemical methods (Baeza et al., 2006); however, these methods have several constraints, such as difficulties regenerating media, chemical wastes, and difficulties to achieve lower than guideline values, especially when the initial concentration of uranium is high (USEPA, 2006). Alternative approaches to uranium remediation utilizing microorganisms are being considered, including biosorption, bioacummulation and biomineralization (Merroun and Selenska-Pobell, 2008). The most accepted biological mechanism for sustainable uranium immobilization in aqueous environments is reductive precipitation, which results from the biologically catalyzed conversion of soluble UO2þ 2 to insoluble UO2(s). Since the first discovery of anaerobic microorganisms capable of reducing U(VI) in the presence of an electron donor (Lovley et al., 1991), a large variety of microorganisms are now known to carry out the reductive precipitation of U(VI) (Merroun and Selenska-Pobell, 2008; Suzuki and Suko, 2006; Wall and Krumholz, 2006). The majority of these microorganisms do not link U(VI)-reduction to energy-gaining processes (Wall and Krumholz, 2006), although there are reports of some Fe(III)-reducing and sulfate-reducing bacteria that can utilize U(VI) as a terminal electron acceptor for energy metabolism (Marshall et al., 2009). Particles of uraninite have been observed to accumulate in the periplasmic space and as deposits occurring externally on the outer membrane around cells (Gorby and Lovley, 1992). Since c-type cytochromes have been localized in zones of U(IV) accumulation, they are hypothesized as required biochemical components for U(VI) bioreduction (Wall and Krumholz, 2006). Reductive biotransformation has been considered as an interesting option for uranium bioremediation at contaminated sites (Lovley and Phillips, 1992). Diverse microbial communities may be stimulated differently depending on the free energies offered by the reaction with different electron donors (Wu et al., 2006). H2 is an example of an effective electron donor for the enzymatic reduction of U(VI) (Marshall
et al., 2009). Ethanol has been also used with success, significantly stimulating the rate of U(VI)-reduction compared to acetate (Luo et al., 2007). Field tests demonstrated that U could be remediated to below EPA’s limit in groundwater using ethanol as electron donor. Reduction of U(VI) to U(IV) in sediments was confirmed (Wu et al., 2007). Anaerobic microbial biofilms can potentially be considered as inoculum to promote U(VI). Anaerobic granular biomass from upward-flow anaerobic sludge blanket (UASB) reactors used for the high-rate treatment of agro-industrial wastewater (Lettinga et al., 1980) is of special interest. The anaerobic granules are highly settleable, have high specific anaerobic activities (Gonzalez-Gil et al., 2001; Lettinga et al., 1980) and possess a high level of biodiversity (Fernandez et al., 2008a). The scope of this work is to determine the extent at which anaerobic granular sludge can perform U(VI) reduction, and investigate its potential application in uranium bioremediation of contaminated groundwater.
2.
Materials and methods
2.1.
Source of biomass
Anaerobic granular biofilms – as the source of inocula – were obtained from full scale upflow anaerobic sludge blanket (UASB) reactors at different wastewater treatment plants. The different granular sludges used were Aviko (Steenderen, The Netherlands), from potato starch processing wastewater (0.115 g volatile suspended solids (VSS) g1 wet wt); Eerbeek (Eerbeek, The Netherlands), from recycled paper wastewater (0.135 g VSS g1 wet wt); Nedalco (Bergen op Zoom, The Netherlands), from a sugar beet distillery effluent (0.0654 g VSS g1 wet wt), and Mahou (Guadalajara, Spain) from beer brewery wastewater (0.0813 g VSS g1 wet wt). The specific acetoclastic methanogenic activity of the sludges were 429, 240, 334 and 408 mg COD g1 VSS d1 for Aviko, Eerbeek, Nedalco and Mahou, respectively. The hydrogenotrophic methanogenic activities measured in Eerbeek and Mahou sludges were 252 and 207 mg COD g1 VSS d1, respectively. All anaerobic granular sludge samples were stored anaerobically at 4 C.
2.2.
Batch experiments
Batch assays were carried out in 160-mL serum bottles (Wheaton, Millville, NJ, USA), containing 100 mL of liquid – mineral basal media – and 60 mL of headspace. The basal media used consisted of 5 mg L1 NH4HCO3, 2 mg L1 K2HPO4, 2.5 mg L1 MgSO4$7H2O, 1 mg L1 Ca(OH)2, 0.33 mg L1 yeast extract, and trace elements in concentration: 0.5 mg L1 H3BO3, 28.0 mg L1 FeSO4$7H2O, 1.06 mg L1 ZnSO4$7H2O, 4.15 mg L1 MnSO4$7H2O, 2.0 mg L1 (NH4)6Mo7O24$4H2O, 1.75 mg L1 AlK (SO4)2$12H2O, 1.13 mg L1 NiSO4$6H2O, 23.6 mg L1 CoSO4$7H2O, 1.0 mg L1 Na2SeO3$5H2O, 5.0 mg L1 Na2WO4$H2O, 1.57 mg L1 CuSO4$5H2O, 10.0 mg L1 EDTA, 2.0 mg L1 resazurin. This media was subsequently adjusted to a pH value of 7.0, and then provided with NaHCO3 to a final concentration of 59 mM. Sodium bicarbonate was used to buffer the pH of the media, as well as a complexing ligand for U(VI), representing the natural
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complexed state of uranium in the groundwater (Elias et al., 2003). U(VI) was provided in the form of uranyl chloride trihydrate (UO2Cl2$3H2O), obtained from International BioAnalytical Industries, Inc. (Boca Raton, FL, USA); 4-mL aliquots from a 10 mM stock solution were added for a final concentration of 0.4 mM U(VI) in each of the bottles. In previous studies, this concentration was determined to be inside the range of non-inhibitory concentrations in terms of biological U(VI)-reducing activity. The granular sludge described above was thoroughly washed in a sieve three times with MilliQ water. The sludge was immediately weighed and transferred to the bottles, which already contained basal media and uranium. The assays were amended with the granular sludge to obtain a final concentration of 0.67 g VSS L1. Several electron donors were added individually as follows: 4.0 mM ethanol and 0.2 mM sodium acetate, providing stoichiometric excesses of 60-fold and 2-fold, respectively based on e equivalents. The bottles were flushed with a gas mixture of N2/CO2 (80:20), first directed to the surface of the liquid with open bottles for 1 min, and then sealed with butyl rubber stoppers and crimp aluminum caps; after this, the N2/CO2 gas mixture was applied for 4 min, this time inserting an inlet and an outlet needle at the top to replace all the remaining oxygen inside the bottle headspace and ensure anaerobic conditions throughout the experimental period. For treatments with H2 as the electron donor, application of N2/CO2 flushing was carried out as described above before H2 was applied. H2 was applied as a gas mixture of H2/CO2 (80:20) with an overpressure of 0.8 atm to sealed bottles by inverting the bottle and direct injection to the liquid phase, in order to provide a final concentration of 19.2 mmol H2 L1 liq , equivalent to a 48fold excess over the stoichiometric requirement of this electron donor based on e equivalents. Controls for the assay were prepared in order to correctly verify the biological and electron donor contribution to the experiment. They consisted of replicated bottles with heatkilled inoculum, as well as bottles without electron donor and without inoculum. In the case of controls with heat-killed inoculum, the corresponding amount of sludge was added to separate bottles with a 10-mL aliquot of MilliQ water 3 days ahead of setting up the experiment, weighed to account for the water losses and covered with aluminum foil. These bottles were autoclaved under the following scheme: an initial sterilization was performed at 121 C for 1 h, allowed to cool down for 24 h; after accomplishing this step, lost water was replaced by weight difference, then autoclaved at the same temperature for 30 min, allowing to cool down again. This last step was repeated the next day. Finally, on the day of the experiment, they were amended with the corresponding amounts of media, uranium and electron donor in order to get 100 mL of liquid. All experiments were performed in duplicated replicates. They were incubated in the dark at 30 C on an orbital shaker 150 rpm. Additionally, the controls described above were incubated along with the treatments. To evaluate the effects of sludge concentration gradients, the same batch set up preparation procedure was followed as described above, except the following concentrations of sludge were used in the different treatments: 0.168, 0.335, 0.67,
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and 1.34 g VSS L1. Abiotic controls (heat-killed inoculum and no inoculum), as well as controls of live inoculum without added electron donor, were prepared accordingly in the concentrations indicated for the live treatments.
2.3.
U(VI) analysis
Liquid samples were taken initially and periodically in subsequent days to measure changes in the soluble uranium concentration. Samples were pippetted into Eppendorf TM centrifuge tubes and immediately centrifuged at 10,000 rpm (RCF of 10,621g) for 10 min. After this step, the supernatant was separated from the tube and transferred to a 3% HNO3 solution. Soluble uranium was measured by using an Inductively Coupled Plasma-Optical Emission Spectrometry (ICP-OES) system model Optima 2100 DV from Perkin–Elmer TM (Shelton, CT, USA). The detection limit for U(VI) was 0.010 mg/L. Since this technique is based on the electromagnetic radiation emission or absorption by an ion in solution, and since U(VI) is being consumed through redox transformation to an insoluble specie U(IV), the removal process was monitored by measuring the intensity of the remaining soluble uranium at a wavelength of 385.958 nm.
2.4.
Respiking of uranyl chloride
After completion of U(VI) consumption in the biologically active treatments, in some experiments fresh 4-mL aliquots of concentrated uranyl chloride (10 mM) were added by injection to the assay bottles, to a final concentration of 0.4 mM U(VI). The bottles were flushed with N2/CO2 (80:20) for 4 min to closed bottles (with an inlet and outlet needle, as described previously), to avoid any possible oxygen contamination inside the bottles. For treatments with H2, the H2/CO2 gas mixture was finally applied with the same procedure as during in the initial feeding to replenish the electron donor. Samples were taken accordingly before and after applying spikes for measurement.
2.5.
Endogenous methane production
In order to account for the level of endogenous electron donor present in Nedalco and Eerbeek sludge, the methane production was measured in each case. This operation consisted in 160-mL bottles amended with 100 mL of basal mineral media and 0.67 g L1 VSS of each type of sludge. A duplicate replicate was made for each of them, and no electron donor was added to the bottles. The bottles were flushed following the procedure previously described for the batch experiments in order to ensure anaerobic conditions in the bottles. Methane gas composition of the headspace was analyzed during the first 30 days of incubation by gas chromatography (GC) using a Hewlett–Packard 5890 Series II system (Agilent Technologies, Palo Alto, CA). It was equipped with a flame ionization detector and a Restek Stabilwax-DA fused silica capillary column (30 m length 0.53 mm ID, Restek Corporation, Bellefonte, PA). Helium was used as the carrier gas at a flow rate of 18 mL min1 and a split flow of 85 mL min1. The
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column was operated at 140 C. The injector port and detector temperatures were 180 and 250 C, respectively.
2.6.
Reoxidation of uraninite
After depletion of uranyl chloride in a batch set of live treatments prepared as described above, a duplicate of treatments was killed by poisoning by adding 1 mL from a 50 mg L1 NaN3 solution along with 0.025 g of HgCl3; other duplicate was left intact (not poisoned) for comparison. Next, a gas mixture consisting of He/CO2/O2 (60:20:20) was flushed into the headspace of the open bottles for 1 min, then 4 min to closed bottles, according to the same flushing procedure used for batch experiments. After this, an overpressure of 1.28 atm (18 psi) of He/CO2/O2 (60:20:20) mixture was applied by inverting bottles and injecting directly to the liquid. This ensured that the concentration was over 6.15 mmol O2 L1 liq .
2.7.
XRD
X-ray diffraction of sludge samples were evaluated to confirm the formation of the U(IV)-containing mineral uraninite. The treatment used for this study was incubated with 0.67 g VSS L1 of Eerbeek sludge and H2 gas as described previously. Uranyl chloride (UO2Cl2$3H2O) was added at an initial concentration of 0.4 mM, and was respiked three times, each spike was added when the previous spike was consumed. After the incubation, the bottles were opened inside an anaerobic chamber (Coy Laboratory Products, Inc.) and solids were carefully separated from the liquid media by decanting the liquid. Solids and remaining liquid media were quickly deposited in a 25-mL vial and sealed with butyl rubber septa and aluminum caps, and were then subjected to drying with ultrapure nitrogen gas. After completely drying, the sample was ground to a fine powder. An X-ray powder diffractometer (Scintag XDS 2000, Cupertino, Ca) was used for the measurement of XRD profiles in the powdered samples. ˚ The following parameters were set: wavelength of 1.5406 A ˚; using Cu-Ka1 radiation; generator settings of 40 kV and 40 mA slits-emitter: 2 mm, 4 mm; receiver: 0.5 mm, 0.3 mm; continuous scan (2q) from 10 to 70 . Raw data were reduced to net intensity by the application of a fast Fourier noise filter, background subtraction and Ka2 stripping. The JCPDS-ICDD database (International Centre for Diffraction Data) was used for the identification of the crystal structures in the sample.
3.
Results
3.1. Intrinsic U(VI) reducing potential of granular anaerobic sludge Batch experiments were performed to test the U(VI) reducing capacity of anaerobic granular sludge from different sources. The U(VI) reducing activity was tested with live and heat-killed inocula with and without hydrogen as electron donor. Fig. 1a and b shows two examples of results obtained for this test. The loss of soluble uranium indicated that the conversion of U(VI) occurred readily in full treatments with live inocula and added H2 gas as electron donor. The removal of U(VI)
Fig. 1 – Time course of the reduction of uranium in the presence of different sources of anaerobic granular sludge: Moderate level of endogenous substrate (A: Eerbeek sludge), and high level of endogenous substrate (B: Nedalco sludge). Legend: –>–, not inoculated; –B–, heat-killed sludge with H2; –:–, live sludge (no electron donor added); and –C–, live sludge with H2.
occurred in the anaerobic sludge without any apparent lag phase. Loss of U(VI) also occurred in the treatments with live inocula and no added electron donor, but the rate of such removal was somewhat lower with one of the sludge inocula (Eerbeek); while mostly not affected with another sludge inoculum (Nedalco). The occurrence of U(VI) removal in the absence of added H2 indicates endogenous substrates in the sludge may have been providing the needed electron donor. No conversion of U(VI) was observed in non-inoculated and heat-killed inoculum controls, even in the presence of added H2 gas, indicating the reactions observed with live inoculum were catalyzed biologically. The rate of removal of U(VI) in the live treatments was for the most part constant indicating zero order kinetics. The two experiments shown in Fig. 1 indicate that the Eerbeek and the Nedalco sludges from different sources had markedly different U(VI) reducing activities with endogenous substrates corresponding to moderate and high rates observed, respectively. However, the rates of removal were similar with added H2 as electron donor, indicating that Nedalco sludge may have had a higher level of endogenous substrate.
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The rate of U(VI) removal was measured in several other samples of anaerobic granular sludge. The results of all the experiments are summarized in Table 1 by showing the average zero order rates in treatments with and without added H2 as electron donor. The table illustrates that of the four different anaerobic sludges tested, only the Eerbeek sludge had a much lower activity (71%) with endogenous substrate when compared with the corresponding treatment with added H2. For the remaining sludge samples, the endogenous rate was only 5–19% lower without an electron donor supplement compared to the treatment with added H2, indicating that the sludge biomass in most of the sludge samples was sufficient to supply the required electron donor. In the presence of added H2, the highest rate was observed with Nedalco sludge, corresponding to a specific activity of 20.5 mg U(VI) g1 VSS d1. However, all of the other sludge inocula had similar specific activities which were only 34–14% lower than the specific activity of the Nedalco sludge.
3.2. Alternative electron donors in the biological removal of U(VI) The potential of alternative electron donors on the U(VI) bioreduction by anaerobic granular sludge was investigated. An experiment was set up to compare reduced organic compounds, acetate and ethanol, with H2 as electron donors in stimulating the activity of U(VI) removal in Eerbeek sludge. The Eerbeek sludge was selected for the experiment since it had the lowest endogenous activity, and thus was considered the most likely sludge to respond to electron donor addition. Fig. 2 shows the concentration of soluble U(VI) decreases readily in all treatments in which live biomass is present. As was observed in the previous experiments, the sludge biomass was observed to have activity in the endogenous treatment. The various electron donors applied varied in their ability to increase the U(VI) removal rate beyond the endogenous rate. A comparison of the removal rates is provided in Table 2. The best stimulation of the rate (2.7 fold) was observed with H2 as electron donor. Ethanol caused a moderate stimulation (of 57%) and acetate had no significant impact on increasing the rate. Removal of uranium was again
Fig. 2 – Comparison of the U(VI) removal achieved in the presence of different electron donors with Eerbeek sludge. –>–, Not inoculated; –-–, heat-killed sludge with acetate; –:–, heat-killed sludge with ethanol; –C–, heat-killed sludge with H2; –B–, live sludge without electron donor added; –-–, live sludge with acetate; –:–, live sludge with ethanol; and –C–, live sludge with H2.
observed to be negligible for the non-inoculated control, as well as for the heat-killed inoculum controls made for each electron donor, confirming that the observed reactions with the live inoculum were biological in nature.
3.3.
Sustainability of U(VI) removal
As part of examining the feasibility of applying anaerobic sludge for the bioreduction of U(VI), it is important to investigate whether the biological reductive activity can be maintained over an extended period of time. To this end, an experiment was set up with several of the sludge inocula to demonstrate sustained removal of U(VI) respiked into the bottles each time the previous allotment of U(VI) was consumed. The experiments were conducted with and without addition of H2 as electron donor. The electron donor (H2) was resupplied after each respiking of U(VI). Abiotic controls form the initial addition of U(VI) were monitored over the entire
Table 1 – Summary of zero order rates obtained from different anaerobic granular biofilm inocula used for the biological reduction of U(VI). Sludge name
Type of wastewater treated
Eerbeeka
Paper recycle
Nedalcoa
Distillery
Aviko
Starch processing
Mahou
Beer brewery
Rate [mM day1]b
Electron donor
H2 None H2 None H2 None H2 None
Average
Standard deviation
Rate Ratio (endog./H2)c
49.4 14.4 57.8 49.7 38.0 35.9 44.9 36.3
4.2 0.5 1.8 2.4 1.4 0.6 0.7 0.4
0.290
a Experiments with this footnote were conducted twice. b Volumetric zero order rate, each bottle contained 0.67 g VSS L1. c Ratio of rate with no added electron donor (endogenous) in numerator over rate with added H2 in denominator.
0.860 0.945 0.809
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Table 2 – Averaged reduction rates for the different electron donors in the presence of the same concentration of Eerbeek sludge. Electron donor added
None Acetate Ethanol Hydrogen
Zero order rate [mM day1]a Average
Standard deviation
16.9 18.5 26.6 45.9
0.6 1.3 3.1 4.2
a Volumetric zero order rate, each bottle contained 0.67 g VSS L1.
course of the experiment to confirm lack of any significant reaction in the absence of biological activity. Fig. 3 shows an example of a respike experiment with the Nedalco sludge. The figure illustrates that the initial high U(VI) reducing activity is maintained for the four consecutive feedings of U(VI) tested. The H2 addition only slightly stimulated the activity, and the degree of stimulation was not greatly changed after respiking U(VI) several times. The abiotic controls remained constant over the course of 103 days, indicating that the reaction observed were biological. These results demonstrate that live sludge sustained its U(VI) removal activity over time under anaerobic conditions, and therefore a continuous system could potentially be applied for the removal of U(VI). In a similar approach, repeated U(VI) spikes were applied to treatments with Eerbeek sludge in the presence and absence of H2 (results not showed). As expected, treatments with added H2 had a higher rate compared to the endogenous treatment. However, both treatments continued to reduce U(VI) throughout the three feedings of U(VI) applied.
3.4.
Evidence of uranium U(IV)
To confirm that U(VI) removal is due to its reduction to insoluble U(IV) an experiment was set up to demonstrate that oxidation by exposure to O2 could resolubilize the uranium. O2 was chosen as the oxidant due to its high standard reduction electron potential compared to other electron acceptors. Fig. 4
Fig. 3 – Time course of the reduction achieved with repeated respikings of U(VI) to Nedalco treatments. –>–, Not inoculated; –C–, heat-killed sludge with H2; –:–, live sludge (no electron donor added); –C–, live sludge with H2.
illustrates that when O2 was introduced at the end of a reduction experiment, the reductively precipitated uranium was readily oxidized and became resolubilized to the original level of U(VI) added. The observation reinforces the hypothesis that U(VI) removal during the anaerobic incubations is due to reduction. In order to determine whether the reoxidation is chemical or biological, some assays with prereduced uranium were poisoned with a mixture of NaN3 and HgCl3. Fig. 4 illustrates that the rate of reoxidation was similar irregardless whether the cells from pre-reduced uranium assays were poisoned or not, indicating that U(IV)-oxidation was due to an abiotic process. XRD spectra of uranium in the sludge also confirmed the presence of U(IV) by the identification of uraninite (UO2) crystals. The XRD pattern of a sample of the sludge solid material after four spikings of 0.4 mM U(VI) and 109 days of incubation with Eerbeek sludge with H2 is shown in Fig. 5. By superimposing the measured diffraction pattern of UO2 (JCPDS-ICDD Card #75-0455), it was observed that the angle 2q of the reflections for UO2 coincide with the position of the peaks in the prepared sample. In addition X-ray photoelectron spectroscopy (XPS) and sequential extraction with water, 1 M bicarbonate and 1 M nitric acid also confirmed the presence of U(VI) as the dominant uranium species in the sample (Supplementary Information).
3.5. Kinetic dependence of U(VI) removal on anaerobic sludge concentration Since anaerobic granular sludge is the likely source of the enzymatic activity for driving the removal of uranium, it should follow that the reducing activity is probably a function of the concentration of sludge present in the treatments. Along this basis, an experiment was carried out by testing different sludge concentrations for the Eerbeek and Nedalco inocula, with moderate and high levels of endogenous
Fig. 4 – Reoxidation of previously bioreduced U(VI) back to U(VI) after addition of O2 gas mixture in treatments (indicated by the dashed vertical line). –>–, Not inoculated; –B–, heat-killed sludge with H2; $$$C$$$, live sludge with H2 and O2 (poisoned); –C–, live sludge with H2 and O2 (nonpoisoned); and –B–, live sludge with H2 (no O2 applied).
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Fig. 5 – XRD pattern of a sample of Eerbeek sludge after reduction of several respikes of uranium during 109 days of incubation. Marks (*) corresponding to the UO2 pattern (JCPDS-ICDD Card #75-0455) are positioned at 2q [ 28.1878, 32.6616, 46.8630, and 55.5873.
substrate, respectively. In both cases, the activity in both the presence and absence of H2 was compared. Fig. 6 shows the time course of the removal achieved at each sludge concentration with the Eerbeek sludge in the presence of H2. The graph shows that as the sludge concentration increase, the removal of U(VI) is more rapid. The initial zero order rates of U(VI) removal are summarized in Fig. 7 as a function of sludge concentration in the presence and absence of added H2. These rates are seen to increase proportionally with the sludge concentration, indicating a dependence of the rate on the sludge concentration. The trend is consistent with the sludge acting as a biocatalyst of the reaction. As expected, in the case of Eerbeek sludge the rates were higher in the presence of the electron donor compared to its absence, even at high sludge concentrations. On the other hand, there was only a small difference in the rates with and without added electron donor for the Nedalco sludge.
Fig. 6 – Removal of U(VI) at different of Eerbeek sludge concentrations. –A–, Not inoculated; heat-killed sludge with H2: –6–, 0.1675 g VSS LL1; –,–, 0.335 g VSS LL1; –>–, 0.67 g VSS LL1; –B–, 1.34 g VSS LL1; live sludge with H2: –:–, 0.1675 g VSS LL1; –-–, 0.335 g VSS LL1; –A–, 0.67 g VSS LL1; –C–, 1.34 g VSS LL1.
Fig. 7 – Correlation of initial zero order rates of U(VI) bioreduction to the anaerobic granular sludge concentration in the presence or absence of added H2 as electron donor. A. Eerbeek sludge, B. Nedalco sludge. –:–, Live sludge with H2; –-–, live sludge without any electron donor added; and –A– killed sludge with H2.
4.
Discussion
4.1. Intrinsic uranium reducing activity of methanogenic sludge In the present study, previously unexposed methanogenic biofilms from UASB reactors readily reduced U(VI). The removal commenced immediately without any observable lag phase. The U(VI) reducing activity could be sustained as evidenced by immediate and rapid removal in experiments when additional spikes of U(VI) were applied. These observations combined with the relatively high specific activities observed (with values ranging from 13.6 to 20.5 mg U(VI) g VSS1 d1) clearly indicate the existence of an initial intrinsic U(VI)reducing capacity in the anaerobic sludge granules. The U(VI)-reducing activity, observed in the presence of unacclimated anaerobic granule sludge, could be explained by the high biodiversity of microorganisms commonly found in methanogenic sludge. The biodiversity includes some microorganisms closely related to known bacteria with the ability to catalyze the reduction of uranium, as indicated by several reviews (Merroun and Selenska-Pobell, 2008; Suzuki and Suko, 2006; Wall and Krumholz, 2006). Clone libraries of 16S rDNA conducted on sludge samples used in this study clearly
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indicate the presence of well-known U(VI)-reducing microorganisms such as Desulfovibrio and Clostrodium spp. in Eerbeek sludge (Fernandez et al., 2008b; Roest et al., 2005). Sequences of 16S rDNA recovered from denature gradient gel electrophoresis bands also indicated the presence of Clostridium in the Nedalco and Mahou sludges (Diaz et al., 2006; Worm et al., 2009). The literature data support the observed intrinsic capacity of anaerobic sludge from different sources to reduce U(VI). The presence of Desulfovibrio, indicates that Eerbeek sludge should have H2-dependent sulfate-reducing activity, which was confirmed by measuring a specific sulfate-reducing 1 VSS d1. activity of 26.35 0.015 mg SO2 4 g
4.2.
Evidence of uranium reduction
To confirm that the mechanism of U(VI) consumption was due to a redox reaction, reoxidation experiments were conducted. This consisted of the addition of oxygen to treatments where the uranium was suspected to have been previously biologically reduced to U(IV). The reduced uranium could be resolubilized by the addition of O2 as oxidant. The oxidation treatment enabled a full recovery of the initial soluble U(VI), suggesting that reductive precipitation was the main mechanism of uranium loss during biological reduction and not a mechanisms such as adsorption. Reoxidation with O2 occurred both in the absence or presence of live (active) cells. The lack of any difference indicates that O2 causes a chemical oxidation of biogenic U(IV). In previous studies, O2 was also shown to readily oxidize previously reduced uranium (Komlos et al., 2008). The evidence that U(IV) was formed was also confirmed by three other methods. XRD demonstrated the presence of U(IV)-containing uraninite crystals in the sludge. XPS confirmed the uranium was in the U(IV) oxidation state. Likewise uranium was in the sludge was not extracted by bicarbonate but was extracted by nitric acid, which is consistent with the presence of U(IV) (Phillips et al., 1995).
4.3.
Effects of the endogenous substrates
In the present study, considerable removal of U(VI) took place in the absence of added electron donors. This fact may be linked to the presence of certain substrates in the anaerobic sludge, which may possibly have electron-donating roles. In previous experiments with sediments, it was observed that removal of uranium occurred without any added electron donor after a certain incubation period (Luo et al., 2007), and it has been suggested that organic matter present in the sediment could be implicated as the electron donor (Marshall et al., 2009). In this study, only a few anaerobic granular sludge samples could be stimulated in a notable way with exogenous electron donors (e.g. Eerbeek), probably because this sludge type had the lowest endogenous electron donor. The level of endogenous electron donor was measured in terms of the methane production over 30 d by Nedalco and Eerbeek sludges. The results showed that the level of endogenous substrate, calculated from the chemical oxygen demand (COD) equivalent of the methane readings after 30 d incubation, was higher in the case of Nedalco (166 mg CH4-COD g VSS1) than that observed for Eerbeek (131 mg CH4-COD g VSS1). The measured
endogenous substrates in these sludges corresponds to 11–14 meq e L1 and only 0.8 meq e L1 is in fact required to reduce the added 0.4 mM U(VI). The reason Eerbeek responded more to the exogenous electron donors (see next heading) is because initially the hydrolysis of the endogenous biomass in that sludge type was very slow. The initial rate of methane release calculated from the readings retrieved over the first 10 days of incubation were 8.9 and 3.4 mg CH4-COD g VSS1 d1 for Nedalco and Eerbeek sludge, respectively. These results suggest that the sludge may contain several different complex organic substrates and degradation intermediates that can contribute in the reduction process, and that the levels of endogenous substrates may vary greatly depending on the origin of the wastewater and the operating conditions of the corresponding UASB system.
4.4. Contribution of the electron donors to the intrinsic activity The presence of an electron donor is an important prerequisite for U(VI) reduction (Wall and Krumholz, 2006). Some of the electron-donating compounds screened in this study (H2, ethanol and acetate) have been found to be effective electron donors in previous U(VI)-bioreduction studies (Anderson and Lovley, 2002). In particular, H2 has been demonstrated to be a very efficient electron donor for U(VI) reduction. Examples of bacteria that readily utilize H2 as an electron donor include Anaeromyxobacter dehalogenans 2CP-C (Marshall et al., 2009), Shewanella spp. (Liu et al., 2002), and Desulfovibrio vulgaris (Lovley et al., 1993). In the present study with the sludge having the lowest endogenous substrate level (Eerbeek), H2 markedly stimulated U(VI)-removal, so that the reaction was complete in only a few days. Even with the other sludges having higher endogenous substrate levels, H2 showed a significant stimulatory effect, albeit that the effect was relatively low. Compared to H2, the other electron donors were less effective. Ethanol had an observable contribution to the removal process in Eerbeek sludge, possibly related to the formation of H2 during the anaerobic degradation of ethanol. Acetate on the other hand, provided no observable stimulation to U(VI) removal beyond the reduction observed with the endogenous substrate. These findings are consistent with previous studies indicating a lower performance of acetate in the reduction of U(VI) compared to H2 (Marshall et al., 2009). Also in studies with ethanol, it has been pointed out that its performance was better than that of acetate (Luo et al., 2007).
4.5. Cell density dependence and its association to the intrinsic activity Reduction of U(VI) occurs only in the presence of live inocula, which suggest a strict enzymatic character of the reaction. The lack of activity in abiotic and heat-killed controls indicate that no other mechanisms can account for U(VI) consumption. Possible alternative mechanisms could have been chemical reduction, sorption to cell material and precipitation of U(VI) with media components such as that known to occur with 3 media containing high levels of NHþ 4 and PO4 (Macaskie et al., 1992; Vazquez et al., 2007). The requirement of the presence of
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live inocula for the uranium reduction to occur has been recognized in previous works (Wall and Krumholz, 2006). Additionally in this study, when the concentration of inocula was increased, the removal rate also increased. This observation confirms the biological nature of the removal process, since this dependence is only seen in the live treatments and not in the controls having the same amount of killed sludge. The increased rates can be attributed to the increase in the enzyme concentration with increasing amounts of active sludge. Mullen et al. (2003) also found that increasing concentrations of Shewanella oneidensis MR-1 resulted in an increased capacity to reduce uranium. Similarly, Spear et al. (1999) found that the lag phases observed in their experiments were inversely proportional to cell concentration. Likewise, they found that by increasing cell concentration, increasing reduction rates were obtained.
5.
Conclusions
The immediate, rapid and sustained U(VI)-removal, observed in this study, suggests that methanogenic anaerobic granular sludge has an innate capacity to support biological removal of U(VI). This phenomenon is most likely due to the natural occurrence of U(VI)-reducing microorganisms in the sludge. The decay of endogenous substrates in anaerobic sludge provides electron equivalents to support U(VI)-reduction. Exogenous electron donors such as H2, stimulates U(VI)removal to varying degrees. A sludge sample with low endogenous substrate levels was stimulated the most with H2 addition. Reoxidation, sequential extraction and spectrometric evidence indicated that the U(VI) removed was converted to insoluble U(IV), confirming that the predominant removal mechanism was reductive precipitation. This study demonstrates the potential feasibility of utilizing granular sludge from UASB reactors for bioremediation of uranium-contaminated groundwater either in ex situ reactors or in permeable reactive barriers.
Acknowledgements This work was supported by the University of Arizona, Technology and Research Initiative Fund (TRIF), Water Sustainability Program. Support was also obtained from the Mexican National Council of Science and Technology (CONACyT ) for a fellowship to A. Tapia-Rodriguez (ref. no. 206827/230752).
Appendix A Supplementary data Supplementary data associated with this article can be found in the online version at, doi:10.1016/j.watres.2009.12.030.
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Available at www.sciencedirect.com
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Anaerobic bioremediation of hexavalent uranium in groundwater by reductive precipitation with methanogenic granular sludge Aida Tapia-Rodriguez, Antonia Luna-Velasco, Jim A. Field*, Reyes Sierra-Alvarez Department of Chemical and Environmental Engineering, University of Arizona, P. O. Box 210011, Tucson, AZ 85721, USA
article info
abstract
Article history:
Uranium has been responsible for extensive contamination of groundwater due to releases
Received 31 August 2009
from mill tailings and other uranium processing waste. Past evidence has confirmed that
Received in revised form
certain bacteria can enzymatically reduce soluble hexavalent uranium (U(VI)) to insoluble
16 December 2009
tetravalent uranium (U(IV)) under anaerobic conditions in the presence of appropriate
Accepted 17 December 2009
electron donors. This paper focuses on the evaluation of anaerobic granular sludge as
Available online 28 December 2009
a source of inoculum for the bioremediation of uranium in water. Batch experiments were performed with several methanogenic anaerobic granular sludge samples and different
Keywords:
electron donors. Abiotic controls consisting of heat-killed inoculum and non-inoculated
UASB
treatments confirmed the biological removal process. In this study, unadapted anaerobic
Uranium reduction
granular sludge immediately reduced U(VI), suggesting an intrinsic capacity of the sludge
Reductive precipitation
to support this process. The high biodiversity of anaerobic granular sludge most likely
Radionuclide
accounts for the presence of specific microorganisms capable of reducing U(VI). Oxidation
Bioreduction
by O2 was shown to resolubilize the uranium. This observation combined with X-ray
Uraninite
diffraction evidence of uraninite confirmed that the removal during anaerobic treatment was due to reductive precipitation. The anaerobic removal activity could be sustained after several respikes of U(VI). The U(VI) removal was feasible without addition of electron donors, indicating that the decay of endogenous biomass substrates was contributing electron equivalents to the process. Addition of electron donors, such as H2 stimulated the removal of U(VI) to varying degrees. The stimulation was greater in sludge samples with lower endogenous substrate levels. The present work reveals the potential application of anaerobic granular sludge for continuous bioremediation schemes to treat uraniumcontaminated water. ª 2009 Elsevier Ltd. All rights reserved.
1.
Introduction
Interest for uranium remediation has increased due to the growing awareness of contamination at mining and processing sites. The environmental contamination results from the
perturbation of naturally occurring uranium minerals through mining and processing in the nuclear fuel cycle as well as during phosphate enrichment. Mill tailings from mining are one of the major causes of extensive uranium contamination due to the large volume of tailings and previous lack of
Abbreviation: VSS, Volatile suspended solids. * Corresponding author. Tel.: þ1 (520) 626 5858; fax: þ1 (520) 621 6048. E-mail address: [email protected] (J.A. Field). 0043-1354/$ – see front matter ª 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2009.12.030
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regulations for their disposal (USEPA, 1995). Uranium is sometimes accompanied by nitrate and sulfate contamination due to the use of nitric and sulfuric acids to extract uranium (Abdelouas, 2006), as well as by the oxidation of primary sulfide minerals from acid mine drainage (USEPA, 1995). The primary public health concern of uranium is its chemical toxicity leading to kidney diseases (ATSDR, 1999). The U.S. Environmental Protection Agency (EPA) has set the primary drinking water limit for uranium at 30 mg L1, which protects the public from kidney toxicity according to the results obtained in epidemiological studies (WHO, 2004). Hexavalent uranium (U(VI)) and tetravalent uranium (U(IV)) are the most common valence states of uranium in nature (Fredrickson et al., 2000). U(VI) predominantly occurs in the uranyl ion form (UO2þ 2 ) (Langmuir, 1978), which is the most reactive and mobile form due to its solubility in water and tendency to form complexes with ligands present in natural waters such as carbonates (Abdelouas, 2006). U(IV) is a highly stable and insoluble form of uranium that generally occurs as the mineral uraninite (UO2(s)). The levels of uranium in the environment are, strongly dependent on pH and redox properties of the subsurface environment. High removal efficiencies (>90%) can be attained by conventional treatment of uranium contamination based on physical-chemical methods (Baeza et al., 2006); however, these methods have several constraints, such as difficulties regenerating media, chemical wastes, and difficulties to achieve lower than guideline values, especially when the initial concentration of uranium is high (USEPA, 2006). Alternative approaches to uranium remediation utilizing microorganisms are being considered, including biosorption, bioacummulation and biomineralization (Merroun and Selenska-Pobell, 2008). The most accepted biological mechanism for sustainable uranium immobilization in aqueous environments is reductive precipitation, which results from the biologically catalyzed conversion of soluble UO2þ 2 to insoluble UO2(s). Since the first discovery of anaerobic microorganisms capable of reducing U(VI) in the presence of an electron donor (Lovley et al., 1991), a large variety of microorganisms are now known to carry out the reductive precipitation of U(VI) (Merroun and Selenska-Pobell, 2008; Suzuki and Suko, 2006; Wall and Krumholz, 2006). The majority of these microorganisms do not link U(VI)-reduction to energy-gaining processes (Wall and Krumholz, 2006), although there are reports of some Fe(III)-reducing and sulfate-reducing bacteria that can utilize U(VI) as a terminal electron acceptor for energy metabolism (Marshall et al., 2009). Particles of uraninite have been observed to accumulate in the periplasmic space and as deposits occurring externally on the outer membrane around cells (Gorby and Lovley, 1992). Since c-type cytochromes have been localized in zones of U(IV) accumulation, they are hypothesized as required biochemical components for U(VI) bioreduction (Wall and Krumholz, 2006). Reductive biotransformation has been considered as an interesting option for uranium bioremediation at contaminated sites (Lovley and Phillips, 1992). Diverse microbial communities may be stimulated differently depending on the free energies offered by the reaction with different electron donors (Wu et al., 2006). H2 is an example of an effective electron donor for the enzymatic reduction of U(VI) (Marshall
et al., 2009). Ethanol has been also used with success, significantly stimulating the rate of U(VI)-reduction compared to acetate (Luo et al., 2007). Field tests demonstrated that U could be remediated to below EPA’s limit in groundwater using ethanol as electron donor. Reduction of U(VI) to U(IV) in sediments was confirmed (Wu et al., 2007). Anaerobic microbial biofilms can potentially be considered as inoculum to promote U(VI). Anaerobic granular biomass from upward-flow anaerobic sludge blanket (UASB) reactors used for the high-rate treatment of agro-industrial wastewater (Lettinga et al., 1980) is of special interest. The anaerobic granules are highly settleable, have high specific anaerobic activities (Gonzalez-Gil et al., 2001; Lettinga et al., 1980) and possess a high level of biodiversity (Fernandez et al., 2008a). The scope of this work is to determine the extent at which anaerobic granular sludge can perform U(VI) reduction, and investigate its potential application in uranium bioremediation of contaminated groundwater.
2.
Materials and methods
2.1.
Source of biomass
Anaerobic granular biofilms – as the source of inocula – were obtained from full scale upflow anaerobic sludge blanket (UASB) reactors at different wastewater treatment plants. The different granular sludges used were Aviko (Steenderen, The Netherlands), from potato starch processing wastewater (0.115 g volatile suspended solids (VSS) g1 wet wt); Eerbeek (Eerbeek, The Netherlands), from recycled paper wastewater (0.135 g VSS g1 wet wt); Nedalco (Bergen op Zoom, The Netherlands), from a sugar beet distillery effluent (0.0654 g VSS g1 wet wt), and Mahou (Guadalajara, Spain) from beer brewery wastewater (0.0813 g VSS g1 wet wt). The specific acetoclastic methanogenic activity of the sludges were 429, 240, 334 and 408 mg COD g1 VSS d1 for Aviko, Eerbeek, Nedalco and Mahou, respectively. The hydrogenotrophic methanogenic activities measured in Eerbeek and Mahou sludges were 252 and 207 mg COD g1 VSS d1, respectively. All anaerobic granular sludge samples were stored anaerobically at 4 C.
2.2.
Batch experiments
Batch assays were carried out in 160-mL serum bottles (Wheaton, Millville, NJ, USA), containing 100 mL of liquid – mineral basal media – and 60 mL of headspace. The basal media used consisted of 5 mg L1 NH4HCO3, 2 mg L1 K2HPO4, 2.5 mg L1 MgSO4$7H2O, 1 mg L1 Ca(OH)2, 0.33 mg L1 yeast extract, and trace elements in concentration: 0.5 mg L1 H3BO3, 28.0 mg L1 FeSO4$7H2O, 1.06 mg L1 ZnSO4$7H2O, 4.15 mg L1 MnSO4$7H2O, 2.0 mg L1 (NH4)6Mo7O24$4H2O, 1.75 mg L1 AlK (SO4)2$12H2O, 1.13 mg L1 NiSO4$6H2O, 23.6 mg L1 CoSO4$7H2O, 1.0 mg L1 Na2SeO3$5H2O, 5.0 mg L1 Na2WO4$H2O, 1.57 mg L1 CuSO4$5H2O, 10.0 mg L1 EDTA, 2.0 mg L1 resazurin. This media was subsequently adjusted to a pH value of 7.0, and then provided with NaHCO3 to a final concentration of 59 mM. Sodium bicarbonate was used to buffer the pH of the media, as well as a complexing ligand for U(VI), representing the natural
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complexed state of uranium in the groundwater (Elias et al., 2003). U(VI) was provided in the form of uranyl chloride trihydrate (UO2Cl2$3H2O), obtained from International BioAnalytical Industries, Inc. (Boca Raton, FL, USA); 4-mL aliquots from a 10 mM stock solution were added for a final concentration of 0.4 mM U(VI) in each of the bottles. In previous studies, this concentration was determined to be inside the range of non-inhibitory concentrations in terms of biological U(VI)-reducing activity. The granular sludge described above was thoroughly washed in a sieve three times with MilliQ water. The sludge was immediately weighed and transferred to the bottles, which already contained basal media and uranium. The assays were amended with the granular sludge to obtain a final concentration of 0.67 g VSS L1. Several electron donors were added individually as follows: 4.0 mM ethanol and 0.2 mM sodium acetate, providing stoichiometric excesses of 60-fold and 2-fold, respectively based on e equivalents. The bottles were flushed with a gas mixture of N2/CO2 (80:20), first directed to the surface of the liquid with open bottles for 1 min, and then sealed with butyl rubber stoppers and crimp aluminum caps; after this, the N2/CO2 gas mixture was applied for 4 min, this time inserting an inlet and an outlet needle at the top to replace all the remaining oxygen inside the bottle headspace and ensure anaerobic conditions throughout the experimental period. For treatments with H2 as the electron donor, application of N2/CO2 flushing was carried out as described above before H2 was applied. H2 was applied as a gas mixture of H2/CO2 (80:20) with an overpressure of 0.8 atm to sealed bottles by inverting the bottle and direct injection to the liquid phase, in order to provide a final concentration of 19.2 mmol H2 L1 liq , equivalent to a 48fold excess over the stoichiometric requirement of this electron donor based on e equivalents. Controls for the assay were prepared in order to correctly verify the biological and electron donor contribution to the experiment. They consisted of replicated bottles with heatkilled inoculum, as well as bottles without electron donor and without inoculum. In the case of controls with heat-killed inoculum, the corresponding amount of sludge was added to separate bottles with a 10-mL aliquot of MilliQ water 3 days ahead of setting up the experiment, weighed to account for the water losses and covered with aluminum foil. These bottles were autoclaved under the following scheme: an initial sterilization was performed at 121 C for 1 h, allowed to cool down for 24 h; after accomplishing this step, lost water was replaced by weight difference, then autoclaved at the same temperature for 30 min, allowing to cool down again. This last step was repeated the next day. Finally, on the day of the experiment, they were amended with the corresponding amounts of media, uranium and electron donor in order to get 100 mL of liquid. All experiments were performed in duplicated replicates. They were incubated in the dark at 30 C on an orbital shaker 150 rpm. Additionally, the controls described above were incubated along with the treatments. To evaluate the effects of sludge concentration gradients, the same batch set up preparation procedure was followed as described above, except the following concentrations of sludge were used in the different treatments: 0.168, 0.335, 0.67,
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and 1.34 g VSS L1. Abiotic controls (heat-killed inoculum and no inoculum), as well as controls of live inoculum without added electron donor, were prepared accordingly in the concentrations indicated for the live treatments.
2.3.
U(VI) analysis
Liquid samples were taken initially and periodically in subsequent days to measure changes in the soluble uranium concentration. Samples were pippetted into Eppendorf TM centrifuge tubes and immediately centrifuged at 10,000 rpm (RCF of 10,621g) for 10 min. After this step, the supernatant was separated from the tube and transferred to a 3% HNO3 solution. Soluble uranium was measured by using an Inductively Coupled Plasma-Optical Emission Spectrometry (ICP-OES) system model Optima 2100 DV from Perkin–Elmer TM (Shelton, CT, USA). The detection limit for U(VI) was 0.010 mg/L. Since this technique is based on the electromagnetic radiation emission or absorption by an ion in solution, and since U(VI) is being consumed through redox transformation to an insoluble specie U(IV), the removal process was monitored by measuring the intensity of the remaining soluble uranium at a wavelength of 385.958 nm.
2.4.
Respiking of uranyl chloride
After completion of U(VI) consumption in the biologically active treatments, in some experiments fresh 4-mL aliquots of concentrated uranyl chloride (10 mM) were added by injection to the assay bottles, to a final concentration of 0.4 mM U(VI). The bottles were flushed with N2/CO2 (80:20) for 4 min to closed bottles (with an inlet and outlet needle, as described previously), to avoid any possible oxygen contamination inside the bottles. For treatments with H2, the H2/CO2 gas mixture was finally applied with the same procedure as during in the initial feeding to replenish the electron donor. Samples were taken accordingly before and after applying spikes for measurement.
2.5.
Endogenous methane production
In order to account for the level of endogenous electron donor present in Nedalco and Eerbeek sludge, the methane production was measured in each case. This operation consisted in 160-mL bottles amended with 100 mL of basal mineral media and 0.67 g L1 VSS of each type of sludge. A duplicate replicate was made for each of them, and no electron donor was added to the bottles. The bottles were flushed following the procedure previously described for the batch experiments in order to ensure anaerobic conditions in the bottles. Methane gas composition of the headspace was analyzed during the first 30 days of incubation by gas chromatography (GC) using a Hewlett–Packard 5890 Series II system (Agilent Technologies, Palo Alto, CA). It was equipped with a flame ionization detector and a Restek Stabilwax-DA fused silica capillary column (30 m length 0.53 mm ID, Restek Corporation, Bellefonte, PA). Helium was used as the carrier gas at a flow rate of 18 mL min1 and a split flow of 85 mL min1. The
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column was operated at 140 C. The injector port and detector temperatures were 180 and 250 C, respectively.
2.6.
Reoxidation of uraninite
After depletion of uranyl chloride in a batch set of live treatments prepared as described above, a duplicate of treatments was killed by poisoning by adding 1 mL from a 50 mg L1 NaN3 solution along with 0.025 g of HgCl3; other duplicate was left intact (not poisoned) for comparison. Next, a gas mixture consisting of He/CO2/O2 (60:20:20) was flushed into the headspace of the open bottles for 1 min, then 4 min to closed bottles, according to the same flushing procedure used for batch experiments. After this, an overpressure of 1.28 atm (18 psi) of He/CO2/O2 (60:20:20) mixture was applied by inverting bottles and injecting directly to the liquid. This ensured that the concentration was over 6.15 mmol O2 L1 liq .
2.7.
XRD
X-ray diffraction of sludge samples were evaluated to confirm the formation of the U(IV)-containing mineral uraninite. The treatment used for this study was incubated with 0.67 g VSS L1 of Eerbeek sludge and H2 gas as described previously. Uranyl chloride (UO2Cl2$3H2O) was added at an initial concentration of 0.4 mM, and was respiked three times, each spike was added when the previous spike was consumed. After the incubation, the bottles were opened inside an anaerobic chamber (Coy Laboratory Products, Inc.) and solids were carefully separated from the liquid media by decanting the liquid. Solids and remaining liquid media were quickly deposited in a 25-mL vial and sealed with butyl rubber septa and aluminum caps, and were then subjected to drying with ultrapure nitrogen gas. After completely drying, the sample was ground to a fine powder. An X-ray powder diffractometer (Scintag XDS 2000, Cupertino, Ca) was used for the measurement of XRD profiles in the powdered samples. ˚ The following parameters were set: wavelength of 1.5406 A ˚; using Cu-Ka1 radiation; generator settings of 40 kV and 40 mA slits-emitter: 2 mm, 4 mm; receiver: 0.5 mm, 0.3 mm; continuous scan (2q) from 10 to 70 . Raw data were reduced to net intensity by the application of a fast Fourier noise filter, background subtraction and Ka2 stripping. The JCPDS-ICDD database (International Centre for Diffraction Data) was used for the identification of the crystal structures in the sample.
3.
Results
3.1. Intrinsic U(VI) reducing potential of granular anaerobic sludge Batch experiments were performed to test the U(VI) reducing capacity of anaerobic granular sludge from different sources. The U(VI) reducing activity was tested with live and heat-killed inocula with and without hydrogen as electron donor. Fig. 1a and b shows two examples of results obtained for this test. The loss of soluble uranium indicated that the conversion of U(VI) occurred readily in full treatments with live inocula and added H2 gas as electron donor. The removal of U(VI)
Fig. 1 – Time course of the reduction of uranium in the presence of different sources of anaerobic granular sludge: Moderate level of endogenous substrate (A: Eerbeek sludge), and high level of endogenous substrate (B: Nedalco sludge). Legend: –>–, not inoculated; –B–, heat-killed sludge with H2; –:–, live sludge (no electron donor added); and –C–, live sludge with H2.
occurred in the anaerobic sludge without any apparent lag phase. Loss of U(VI) also occurred in the treatments with live inocula and no added electron donor, but the rate of such removal was somewhat lower with one of the sludge inocula (Eerbeek); while mostly not affected with another sludge inoculum (Nedalco). The occurrence of U(VI) removal in the absence of added H2 indicates endogenous substrates in the sludge may have been providing the needed electron donor. No conversion of U(VI) was observed in non-inoculated and heat-killed inoculum controls, even in the presence of added H2 gas, indicating the reactions observed with live inoculum were catalyzed biologically. The rate of removal of U(VI) in the live treatments was for the most part constant indicating zero order kinetics. The two experiments shown in Fig. 1 indicate that the Eerbeek and the Nedalco sludges from different sources had markedly different U(VI) reducing activities with endogenous substrates corresponding to moderate and high rates observed, respectively. However, the rates of removal were similar with added H2 as electron donor, indicating that Nedalco sludge may have had a higher level of endogenous substrate.
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The rate of U(VI) removal was measured in several other samples of anaerobic granular sludge. The results of all the experiments are summarized in Table 1 by showing the average zero order rates in treatments with and without added H2 as electron donor. The table illustrates that of the four different anaerobic sludges tested, only the Eerbeek sludge had a much lower activity (71%) with endogenous substrate when compared with the corresponding treatment with added H2. For the remaining sludge samples, the endogenous rate was only 5–19% lower without an electron donor supplement compared to the treatment with added H2, indicating that the sludge biomass in most of the sludge samples was sufficient to supply the required electron donor. In the presence of added H2, the highest rate was observed with Nedalco sludge, corresponding to a specific activity of 20.5 mg U(VI) g1 VSS d1. However, all of the other sludge inocula had similar specific activities which were only 34–14% lower than the specific activity of the Nedalco sludge.
3.2. Alternative electron donors in the biological removal of U(VI) The potential of alternative electron donors on the U(VI) bioreduction by anaerobic granular sludge was investigated. An experiment was set up to compare reduced organic compounds, acetate and ethanol, with H2 as electron donors in stimulating the activity of U(VI) removal in Eerbeek sludge. The Eerbeek sludge was selected for the experiment since it had the lowest endogenous activity, and thus was considered the most likely sludge to respond to electron donor addition. Fig. 2 shows the concentration of soluble U(VI) decreases readily in all treatments in which live biomass is present. As was observed in the previous experiments, the sludge biomass was observed to have activity in the endogenous treatment. The various electron donors applied varied in their ability to increase the U(VI) removal rate beyond the endogenous rate. A comparison of the removal rates is provided in Table 2. The best stimulation of the rate (2.7 fold) was observed with H2 as electron donor. Ethanol caused a moderate stimulation (of 57%) and acetate had no significant impact on increasing the rate. Removal of uranium was again
Fig. 2 – Comparison of the U(VI) removal achieved in the presence of different electron donors with Eerbeek sludge. –>–, Not inoculated; –-–, heat-killed sludge with acetate; –:–, heat-killed sludge with ethanol; –C–, heat-killed sludge with H2; –B–, live sludge without electron donor added; –-–, live sludge with acetate; –:–, live sludge with ethanol; and –C–, live sludge with H2.
observed to be negligible for the non-inoculated control, as well as for the heat-killed inoculum controls made for each electron donor, confirming that the observed reactions with the live inoculum were biological in nature.
3.3.
Sustainability of U(VI) removal
As part of examining the feasibility of applying anaerobic sludge for the bioreduction of U(VI), it is important to investigate whether the biological reductive activity can be maintained over an extended period of time. To this end, an experiment was set up with several of the sludge inocula to demonstrate sustained removal of U(VI) respiked into the bottles each time the previous allotment of U(VI) was consumed. The experiments were conducted with and without addition of H2 as electron donor. The electron donor (H2) was resupplied after each respiking of U(VI). Abiotic controls form the initial addition of U(VI) were monitored over the entire
Table 1 – Summary of zero order rates obtained from different anaerobic granular biofilm inocula used for the biological reduction of U(VI). Sludge name
Type of wastewater treated
Eerbeeka
Paper recycle
Nedalcoa
Distillery
Aviko
Starch processing
Mahou
Beer brewery
Rate [mM day1]b
Electron donor
H2 None H2 None H2 None H2 None
Average
Standard deviation
Rate Ratio (endog./H2)c
49.4 14.4 57.8 49.7 38.0 35.9 44.9 36.3
4.2 0.5 1.8 2.4 1.4 0.6 0.7 0.4
0.290
a Experiments with this footnote were conducted twice. b Volumetric zero order rate, each bottle contained 0.67 g VSS L1. c Ratio of rate with no added electron donor (endogenous) in numerator over rate with added H2 in denominator.
0.860 0.945 0.809
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Table 2 – Averaged reduction rates for the different electron donors in the presence of the same concentration of Eerbeek sludge. Electron donor added
None Acetate Ethanol Hydrogen
Zero order rate [mM day1]a Average
Standard deviation
16.9 18.5 26.6 45.9
0.6 1.3 3.1 4.2
a Volumetric zero order rate, each bottle contained 0.67 g VSS L1.
course of the experiment to confirm lack of any significant reaction in the absence of biological activity. Fig. 3 shows an example of a respike experiment with the Nedalco sludge. The figure illustrates that the initial high U(VI) reducing activity is maintained for the four consecutive feedings of U(VI) tested. The H2 addition only slightly stimulated the activity, and the degree of stimulation was not greatly changed after respiking U(VI) several times. The abiotic controls remained constant over the course of 103 days, indicating that the reaction observed were biological. These results demonstrate that live sludge sustained its U(VI) removal activity over time under anaerobic conditions, and therefore a continuous system could potentially be applied for the removal of U(VI). In a similar approach, repeated U(VI) spikes were applied to treatments with Eerbeek sludge in the presence and absence of H2 (results not showed). As expected, treatments with added H2 had a higher rate compared to the endogenous treatment. However, both treatments continued to reduce U(VI) throughout the three feedings of U(VI) applied.
3.4.
Evidence of uranium U(IV)
To confirm that U(VI) removal is due to its reduction to insoluble U(IV) an experiment was set up to demonstrate that oxidation by exposure to O2 could resolubilize the uranium. O2 was chosen as the oxidant due to its high standard reduction electron potential compared to other electron acceptors. Fig. 4
Fig. 3 – Time course of the reduction achieved with repeated respikings of U(VI) to Nedalco treatments. –>–, Not inoculated; –C–, heat-killed sludge with H2; –:–, live sludge (no electron donor added); –C–, live sludge with H2.
illustrates that when O2 was introduced at the end of a reduction experiment, the reductively precipitated uranium was readily oxidized and became resolubilized to the original level of U(VI) added. The observation reinforces the hypothesis that U(VI) removal during the anaerobic incubations is due to reduction. In order to determine whether the reoxidation is chemical or biological, some assays with prereduced uranium were poisoned with a mixture of NaN3 and HgCl3. Fig. 4 illustrates that the rate of reoxidation was similar irregardless whether the cells from pre-reduced uranium assays were poisoned or not, indicating that U(IV)-oxidation was due to an abiotic process. XRD spectra of uranium in the sludge also confirmed the presence of U(IV) by the identification of uraninite (UO2) crystals. The XRD pattern of a sample of the sludge solid material after four spikings of 0.4 mM U(VI) and 109 days of incubation with Eerbeek sludge with H2 is shown in Fig. 5. By superimposing the measured diffraction pattern of UO2 (JCPDS-ICDD Card #75-0455), it was observed that the angle 2q of the reflections for UO2 coincide with the position of the peaks in the prepared sample. In addition X-ray photoelectron spectroscopy (XPS) and sequential extraction with water, 1 M bicarbonate and 1 M nitric acid also confirmed the presence of U(VI) as the dominant uranium species in the sample (Supplementary Information).
3.5. Kinetic dependence of U(VI) removal on anaerobic sludge concentration Since anaerobic granular sludge is the likely source of the enzymatic activity for driving the removal of uranium, it should follow that the reducing activity is probably a function of the concentration of sludge present in the treatments. Along this basis, an experiment was carried out by testing different sludge concentrations for the Eerbeek and Nedalco inocula, with moderate and high levels of endogenous
Fig. 4 – Reoxidation of previously bioreduced U(VI) back to U(VI) after addition of O2 gas mixture in treatments (indicated by the dashed vertical line). –>–, Not inoculated; –B–, heat-killed sludge with H2; $$$C$$$, live sludge with H2 and O2 (poisoned); –C–, live sludge with H2 and O2 (nonpoisoned); and –B–, live sludge with H2 (no O2 applied).
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Fig. 5 – XRD pattern of a sample of Eerbeek sludge after reduction of several respikes of uranium during 109 days of incubation. Marks (*) corresponding to the UO2 pattern (JCPDS-ICDD Card #75-0455) are positioned at 2q [ 28.1878, 32.6616, 46.8630, and 55.5873.
substrate, respectively. In both cases, the activity in both the presence and absence of H2 was compared. Fig. 6 shows the time course of the removal achieved at each sludge concentration with the Eerbeek sludge in the presence of H2. The graph shows that as the sludge concentration increase, the removal of U(VI) is more rapid. The initial zero order rates of U(VI) removal are summarized in Fig. 7 as a function of sludge concentration in the presence and absence of added H2. These rates are seen to increase proportionally with the sludge concentration, indicating a dependence of the rate on the sludge concentration. The trend is consistent with the sludge acting as a biocatalyst of the reaction. As expected, in the case of Eerbeek sludge the rates were higher in the presence of the electron donor compared to its absence, even at high sludge concentrations. On the other hand, there was only a small difference in the rates with and without added electron donor for the Nedalco sludge.
Fig. 6 – Removal of U(VI) at different of Eerbeek sludge concentrations. –A–, Not inoculated; heat-killed sludge with H2: –6–, 0.1675 g VSS LL1; –,–, 0.335 g VSS LL1; –>–, 0.67 g VSS LL1; –B–, 1.34 g VSS LL1; live sludge with H2: –:–, 0.1675 g VSS LL1; –-–, 0.335 g VSS LL1; –A–, 0.67 g VSS LL1; –C–, 1.34 g VSS LL1.
Fig. 7 – Correlation of initial zero order rates of U(VI) bioreduction to the anaerobic granular sludge concentration in the presence or absence of added H2 as electron donor. A. Eerbeek sludge, B. Nedalco sludge. –:–, Live sludge with H2; –-–, live sludge without any electron donor added; and –A– killed sludge with H2.
4.
Discussion
4.1. Intrinsic uranium reducing activity of methanogenic sludge In the present study, previously unexposed methanogenic biofilms from UASB reactors readily reduced U(VI). The removal commenced immediately without any observable lag phase. The U(VI) reducing activity could be sustained as evidenced by immediate and rapid removal in experiments when additional spikes of U(VI) were applied. These observations combined with the relatively high specific activities observed (with values ranging from 13.6 to 20.5 mg U(VI) g VSS1 d1) clearly indicate the existence of an initial intrinsic U(VI)reducing capacity in the anaerobic sludge granules. The U(VI)-reducing activity, observed in the presence of unacclimated anaerobic granule sludge, could be explained by the high biodiversity of microorganisms commonly found in methanogenic sludge. The biodiversity includes some microorganisms closely related to known bacteria with the ability to catalyze the reduction of uranium, as indicated by several reviews (Merroun and Selenska-Pobell, 2008; Suzuki and Suko, 2006; Wall and Krumholz, 2006). Clone libraries of 16S rDNA conducted on sludge samples used in this study clearly
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indicate the presence of well-known U(VI)-reducing microorganisms such as Desulfovibrio and Clostrodium spp. in Eerbeek sludge (Fernandez et al., 2008b; Roest et al., 2005). Sequences of 16S rDNA recovered from denature gradient gel electrophoresis bands also indicated the presence of Clostridium in the Nedalco and Mahou sludges (Diaz et al., 2006; Worm et al., 2009). The literature data support the observed intrinsic capacity of anaerobic sludge from different sources to reduce U(VI). The presence of Desulfovibrio, indicates that Eerbeek sludge should have H2-dependent sulfate-reducing activity, which was confirmed by measuring a specific sulfate-reducing 1 VSS d1. activity of 26.35 0.015 mg SO2 4 g
4.2.
Evidence of uranium reduction
To confirm that the mechanism of U(VI) consumption was due to a redox reaction, reoxidation experiments were conducted. This consisted of the addition of oxygen to treatments where the uranium was suspected to have been previously biologically reduced to U(IV). The reduced uranium could be resolubilized by the addition of O2 as oxidant. The oxidation treatment enabled a full recovery of the initial soluble U(VI), suggesting that reductive precipitation was the main mechanism of uranium loss during biological reduction and not a mechanisms such as adsorption. Reoxidation with O2 occurred both in the absence or presence of live (active) cells. The lack of any difference indicates that O2 causes a chemical oxidation of biogenic U(IV). In previous studies, O2 was also shown to readily oxidize previously reduced uranium (Komlos et al., 2008). The evidence that U(IV) was formed was also confirmed by three other methods. XRD demonstrated the presence of U(IV)-containing uraninite crystals in the sludge. XPS confirmed the uranium was in the U(IV) oxidation state. Likewise uranium was in the sludge was not extracted by bicarbonate but was extracted by nitric acid, which is consistent with the presence of U(IV) (Phillips et al., 1995).
4.3.
Effects of the endogenous substrates
In the present study, considerable removal of U(VI) took place in the absence of added electron donors. This fact may be linked to the presence of certain substrates in the anaerobic sludge, which may possibly have electron-donating roles. In previous experiments with sediments, it was observed that removal of uranium occurred without any added electron donor after a certain incubation period (Luo et al., 2007), and it has been suggested that organic matter present in the sediment could be implicated as the electron donor (Marshall et al., 2009). In this study, only a few anaerobic granular sludge samples could be stimulated in a notable way with exogenous electron donors (e.g. Eerbeek), probably because this sludge type had the lowest endogenous electron donor. The level of endogenous electron donor was measured in terms of the methane production over 30 d by Nedalco and Eerbeek sludges. The results showed that the level of endogenous substrate, calculated from the chemical oxygen demand (COD) equivalent of the methane readings after 30 d incubation, was higher in the case of Nedalco (166 mg CH4-COD g VSS1) than that observed for Eerbeek (131 mg CH4-COD g VSS1). The measured
endogenous substrates in these sludges corresponds to 11–14 meq e L1 and only 0.8 meq e L1 is in fact required to reduce the added 0.4 mM U(VI). The reason Eerbeek responded more to the exogenous electron donors (see next heading) is because initially the hydrolysis of the endogenous biomass in that sludge type was very slow. The initial rate of methane release calculated from the readings retrieved over the first 10 days of incubation were 8.9 and 3.4 mg CH4-COD g VSS1 d1 for Nedalco and Eerbeek sludge, respectively. These results suggest that the sludge may contain several different complex organic substrates and degradation intermediates that can contribute in the reduction process, and that the levels of endogenous substrates may vary greatly depending on the origin of the wastewater and the operating conditions of the corresponding UASB system.
4.4. Contribution of the electron donors to the intrinsic activity The presence of an electron donor is an important prerequisite for U(VI) reduction (Wall and Krumholz, 2006). Some of the electron-donating compounds screened in this study (H2, ethanol and acetate) have been found to be effective electron donors in previous U(VI)-bioreduction studies (Anderson and Lovley, 2002). In particular, H2 has been demonstrated to be a very efficient electron donor for U(VI) reduction. Examples of bacteria that readily utilize H2 as an electron donor include Anaeromyxobacter dehalogenans 2CP-C (Marshall et al., 2009), Shewanella spp. (Liu et al., 2002), and Desulfovibrio vulgaris (Lovley et al., 1993). In the present study with the sludge having the lowest endogenous substrate level (Eerbeek), H2 markedly stimulated U(VI)-removal, so that the reaction was complete in only a few days. Even with the other sludges having higher endogenous substrate levels, H2 showed a significant stimulatory effect, albeit that the effect was relatively low. Compared to H2, the other electron donors were less effective. Ethanol had an observable contribution to the removal process in Eerbeek sludge, possibly related to the formation of H2 during the anaerobic degradation of ethanol. Acetate on the other hand, provided no observable stimulation to U(VI) removal beyond the reduction observed with the endogenous substrate. These findings are consistent with previous studies indicating a lower performance of acetate in the reduction of U(VI) compared to H2 (Marshall et al., 2009). Also in studies with ethanol, it has been pointed out that its performance was better than that of acetate (Luo et al., 2007).
4.5. Cell density dependence and its association to the intrinsic activity Reduction of U(VI) occurs only in the presence of live inocula, which suggest a strict enzymatic character of the reaction. The lack of activity in abiotic and heat-killed controls indicate that no other mechanisms can account for U(VI) consumption. Possible alternative mechanisms could have been chemical reduction, sorption to cell material and precipitation of U(VI) with media components such as that known to occur with 3 media containing high levels of NHþ 4 and PO4 (Macaskie et al., 1992; Vazquez et al., 2007). The requirement of the presence of
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live inocula for the uranium reduction to occur has been recognized in previous works (Wall and Krumholz, 2006). Additionally in this study, when the concentration of inocula was increased, the removal rate also increased. This observation confirms the biological nature of the removal process, since this dependence is only seen in the live treatments and not in the controls having the same amount of killed sludge. The increased rates can be attributed to the increase in the enzyme concentration with increasing amounts of active sludge. Mullen et al. (2003) also found that increasing concentrations of Shewanella oneidensis MR-1 resulted in an increased capacity to reduce uranium. Similarly, Spear et al. (1999) found that the lag phases observed in their experiments were inversely proportional to cell concentration. Likewise, they found that by increasing cell concentration, increasing reduction rates were obtained.
5.
Conclusions
The immediate, rapid and sustained U(VI)-removal, observed in this study, suggests that methanogenic anaerobic granular sludge has an innate capacity to support biological removal of U(VI). This phenomenon is most likely due to the natural occurrence of U(VI)-reducing microorganisms in the sludge. The decay of endogenous substrates in anaerobic sludge provides electron equivalents to support U(VI)-reduction. Exogenous electron donors such as H2, stimulates U(VI)removal to varying degrees. A sludge sample with low endogenous substrate levels was stimulated the most with H2 addition. Reoxidation, sequential extraction and spectrometric evidence indicated that the U(VI) removed was converted to insoluble U(IV), confirming that the predominant removal mechanism was reductive precipitation. This study demonstrates the potential feasibility of utilizing granular sludge from UASB reactors for bioremediation of uranium-contaminated groundwater either in ex situ reactors or in permeable reactive barriers.
Acknowledgements This work was supported by the University of Arizona, Technology and Research Initiative Fund (TRIF), Water Sustainability Program. Support was also obtained from the Mexican National Council of Science and Technology (CONACyT ) for a fellowship to A. Tapia-Rodriguez (ref. no. 206827/230752).
Appendix A Supplementary data Supplementary data associated with this article can be found in the online version at, doi:10.1016/j.watres.2009.12.030.
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