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Published online 13 September 2006
Published in J Environ Qual 35:1873-1883 (2006)
DOI: 10.2134/jeq2005.0443
© 2006 American Society of Agronomy, Crop Science Society of America, and Soil Science Society of America
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TECHNICAL REPORTS

Heavy Metals in the Environment

Bioconversion of Selenate in Methanogenic Anaerobic Granular Sludge

Violeta Astratineib, Eric van Hullebuschc and Piet Lensa,*

a Sub-department of Environmental Technology, Wageningen University, P.O. Box 8129, EV Wageningen 6700, The Netherlands
b National Research and Development Institute for Environmental Protection, Splaiul Independentei 294, Bucharest 77703, Romania
c Laboratoire des Géomatériaux et Géologie de l'Ingénieur; Université de Marne la Vallée; Institut Francilien des Sciences Appliquées, Bât. IFI; 5, Boulevard Descartes-Champs sur Marne; 77454 Marne La Vallée, Cedex 2, France

* Corresponding author (piet.lens{at}wur.nl)

Received for publication November 30, 2005.

   ABSTRACT
TOP
 ABSTRACT
INTRODUCTION
 MATERIAL AND METHODS
 RESULTS
 DISCUSSION
 REFERENCES
 
The capacity of anaerobic granular sludge to remove selenate from contaminated wastewater was investigated. The potential of different types of granular sludge to remove selenate from the liquid phase was compared to that of suspended sludge and contaminated soil and sediment samples. The selenate removal rates ranged from 400 to 1500 µg g VSS–1 h–1, depending on the source of biomass, electron donor, and the initial selenate concentration. The granular structure protects the microorganisms when exposed to high selenate concentrations (0.1 to 1 mM). Anaerobic granular sludge "Eerbeek," originating from a UASB reactor treating paper mill wastewater, removed about 90, 50, and 36% of 0.1, 0.5, and 1 mM of Se, respectively, from the liquid phase when incubated with 20 mM lactate at 30°C and pH 7.5. Selenite, elemental Se (Seo), and metal selenide precipitates were the conversion products. Enrichments from the anaerobic granular sludge "Eerbeek" were able to convert 90% of the 10-mM selenate to Seo at a rate of 1505 µg Se(VI) g cells–1 h–1, a specific growth rate of 0.0125 g cells h–1, and a yield of 0.083 g cells mg Se–1. Both microbial metabolic processes (e.g dissimilatory reduction) as well as microbially mediated physicochemical mechanisms (adsorption and precipitation) contribute to the removal of selenate from the Se-containing medium.

Abbreviations: VSS, volatile suspended solids • Seo, elemental selenium


   INTRODUCTION
TOP
 ABSTRACT
 INTRODUCTION
MATERIAL AND METHODS
 RESULTS
 DISCUSSION
 REFERENCES
 
THE TOXICITY AND BIOAVAILABILITY of Se largely depends on the chemical oxidation state of the element (Amweg et al., 2002). Selenium is present in the environment in four different oxidation states (-II, 0, IV and VI) as a variety of inorganic (selenate, selenite, Seo, selenides) and organic (amino-acids and methylated) compounds (Linkson, 1992; Fan Teresa et al., 2002). Selenate (SeO42–) and selenite (SeO32–) are both water soluble species and possess toxic effects to organisms at high concentrations (Wu, 2004). Elemental Se has been commonly considered an unavailable form of Se because of its insolubility. Hydrogen selenide (HSe–) is a highly toxic gas, which is spontaneously and rapidly oxidized to non-toxic Seo in the presence of air (Herbel et al., 2003). However, under strongly reducing conditions, selenides [Se(-II)] are thermodynamically stable and form insoluble metal selenide precipitates (Séby et al., 1998).

Selenium contamination of wastewater is related to industrial activities (Lemly, 2004). For example, wastewater from coal mining (3 to 330 µg Se L–1), gold mining (200 to 33 000 µg Se L–1), and fuel refining (170 to 4900 µg Se L–1) contain elevated levels of soluble Se as selenate and selenite (Twidwell et al., 1999). Also, drainage or irrigation in naturally Se-rich areas can lead to elevated selenate (up to 300 µg L–1) concentrations in water (Macy et al., 1993; Oremland et al., 1999; de Souza et al., 2001; Wu, 2004). Chronic poisoning of animal populations with Se can result in lethality, reproductive malformations, or typical illnesses known as "alkali disease" and "blind staggers" (Tinggi, 2003; Hamilton, 2004). Human intake of soluble Se produces chronic selenosis with dermal and neurologic effects (Yan et al., 1983; USEPA, 2003). Recently, it has been found that bioavailable Se compounds as selenite and selenate can be selectively toxic to human cells, inducing necrosis (Weiller et al., 2004). Therefore, cost-effective removal technologies for Se oxyanions need to be developed.

Physicochemical methods such as chemical precipitation, catalytic reduction with zero valent iron, ion exchange, and membrane separation technologies, mainly applied for removing Se oxyanions from wastewater (Kashiva et al., 2000; Fujita et al., 2002), are expensive and often result in secondary effects in the environment (Kessi et al., 1999). Several biological conversion routes of oxidized Se compounds have also been proposed for Se removal from surface and groundwater. These include assimilatory reduction (Turner et al., 1998), dissimilatory reduction (Stolz and Oremland, 1999; Herbel et al., 2000; Stolz et al., 2002), and biomethylation (van Fleet-Stalder et al., 2000). These biological processes for the removal of soluble Se were mainly described for different natural inocula (soil and sediments) from which most of the bacteria able to reduce selenate or selenite were isolated (Ike et al., 2000; de Souza et al., 2001). To the best of our knowledge, the mechanism and capacity of biological selenate removal of anaerobic methanogenic granular sludge has so far not been investigated. The performance of anaerobic granular sludge was described previously for the degradation of complex organic wastewaters (Lettinga, 1995), as well as the removal of nitrate (van der Maas et al., 2003) or sulfate (Lens et al., 1999). The functional diversity of microbial populations, i.e., hydrolytic, fermentative, acidogenic, homoacetogenic, acetogenic, sulfate-reducing, and methanogenic bacteria (Santegoeds et al., 1999) of granular sludge was found to play an important role in the microbial removal of different types of organic and inorganic pollutants from wastewaters.

This paper reports on the capacity of methanogenic granular sludge for selenate (SeO42–) reduction at pH 7.5 and 30°C in comparison to that of inocula from other environments. The influence of biomass crushing, different selenate concentrations and the type of electron donor on the SeO42– removal rate was investigated in batch experiments. In addition, studies on the SeO42– conversion mechanism by anaerobic granular biomass were performed.


   MATERIAL AND METHODS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIAL AND METHODS
RESULTS
 DISCUSSION
 REFERENCES
 
Experimental Setup
Experiments were performed in 250-mL serum bottles, filled with 100 mL basal mineral medium containing different selenate concentrations and around 2 g wet anaerobic granular sludge, sediment, or soil. The basal mineral medium contained (g L–1 demiwater): NH4Cl (0.3), NaCl (0.3), CaCl2 2H2O (0.11), MgCl2 6H2O (0.1), and 1 mL L–1 trace element solution according to Stams et al. (1993). Unless stated otherwise, 0.5 g L–1 of yeast extract was supplemented to the basal mineral medium. The medium was buffered with 1.14 g L–1 KH2PO4 and 2.03 g L–1 Na2HPO4 2H2O. The reducing agent was Na2S (0.288 g L–1 final concentration) and the redox indicator resazurine (0.0005 g L–1 final concentration). The pH was adjusted to 7.5 by adding NaOH. Selenate was added from concentrated stock solutions to achieve initial concentrations of 0.1 mM, 0.5 mM, 1 mM, 10 mM, and 20 mM. Experiments were performed at least in duplicate.

Lactate was added to the basal mineral medium as sodium lactate solution (50%) to obtain an initial concentration of 20 mM. To investigate the effect of the electron donor on selenate reduction by Eerbeek sludge, acetate, propionate, glucose, sucrose, casein hydrolisate, ethanol, and methanol (initial concentration of 20 mM) were used as well. Controls without any substrate addition were incubated under the same conditions.

Tests were performed with intact granules, unless specified otherwise. If required, granular sludge was crushed by pushing the granules by a syringe through needles with descending diameter under anaerobic conditions, as described by Sipma et al. (2003). Controls were incubated with sterilized (autoclaved for 20 min at 121°C) sludge (Eerbeek and Nedalco) and without sludge for abiotic selenate conversions.

After closing the bottles with butyl rubber septa (Rubber B.V., Hilversum, The Netherlands) and aluminum caps, the headspace was flushed with an excess of oxygen-free N2 for 5 min (final pressure 1.78 bar). Serum bottles were incubated in the dark at 30°C and shaken at 85 rpm.

Source of Biomass
Experiments were performed with biomass sampled from full scale wastewater treatment plants (Table 1): methanogenic granular sludge from a full scale UASB reactor treating paper mill wastewater (Eerbeek, The Netherlands), methanogenic granular sludge from a full scale UASB reactor treating alcohol distillery wastewater (Nedalco, Bergen op Zoom, The Netherlands), denitrifying sludge from a fluidized bed reactor treating nitrate-polluted surface water (Veendam, The Netherlands), and selenogenic suspended sludge from a 1 m3 h pilot plant at Umicore (Hoboken, Belgium) treating Se-contaminated wastewater (5 mg.L–1) from photographic, electronic, and catalytic applications. Soil and river sediment from a mining area (Baia Mare, Romania) contaminated with heavy metals (Cu, Cr, Fe, Cd, Zn, As, and Se) were included to compare their selenate reduction potential with that of wastewater treatment sludges. Sediment samples from the Wadden Sea (The Netherlands) were used as well.


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  		Table 1. Source of inoculum.

 
Enrichment of Selenate-Reducing Bacteria
Selenate removal tests were performed with an enrichment of selenate-reducing bacteria for studying their kinetic parameters in comparison with that of granular sludge. For this, aliquots of supernatant from batch bottles with Eerbeek sludge were inoculated in a selenate- (10 mM) containing mineral basal medium supplemented with 20 mM lactate and cultivated for 10 to 14 d under anaerobic conditions at 30°C and pH 7.5. The resultant culture was three times transferred (1.5 mL) into a new medium under the same conditions for the enrichment and maintenance of selenate-reducing bacteria.

Growth was monitored by determining the absorbance at 600 nm (Garbisu et al., 1995) by a Spectronics 60 Milton-Roy spectrophotometer (Analytical Products Division, Belgium). The absorbance values of the enrichment cultures were correlated by regression analysis with the mass (dry weight) of the cells (g cell L–1) present in 5 mL cultures obtained after filtering and overnight drying, according to Caroll (1999). The kinetic parameters reported are the average of the values obtained in two transfers.

Analytical Methods
Ethanol, methanol, acetate, and propionate were measured by gas chromatography (Hewlett-Packard HP 5890, Palo Alto, USA) as described by Weijma et al. (2000). Lactate was analyzed by high pressure liquid chromatography as described in Gonzalez-Gil et al. (2003). The volatile suspended solids (VSS) content was determined at the end of the tests as described by Standard Methods (APHA, 1995). Soluble Se (selenate, selenite) in the supernatant fraction was measured by ion chromatography (DX-600 IC system, Dionex, Salt Lake City, USA) after 5 min centrifugation of the samples at 16.760 g. All chemicals were of analytical or biological grade and purchased from E. Merck AG (Darmstadt, Germany).

Quantitative analysis for selenium was performed using an inductively coupled plasma-optical emission spectrometer (ICP-OES) (Varian-vista, Australia) equipped with an ultrasonic nebulizer (USN) for direct aqueous sample analysis at µg L–1 level. For total Se measurements in the liquid phase, hydrogen peroxide (H2O2) was added until a colorless liquid was obtained (Profumo et al., 2001). An optimal analytical wavelength of 196.026 nm was chosen to perform the analysis. The detection limit was 6 µg L–1; the relative standard deviation (RSD) for the Se content in the samples was in the 0.6 to 3.6% range. Each sample was analyzed in triplicate.

The pseudo-total metal content of the sludge was determined by digestion with the aqua regia procedure. Total suspended solids (TSS) of 0.5 g of anaerobic granular sludge was treated with 10 mL of aqua regia in TEFLON PFA (PerFluoroAlkoxy resin) digestion vessels in a temperature-controlled microwave oven (Milestone ETHOS E, Milestone; Monroe, CT). The digests were filtered through a Schleicher and Schuell 589 filter paper into volumetric flasks and then completed to 100 mL with ultra pure water for quantitative Se analysis (ICP-OES). Low-magnification microscopy and light microscopy were done using an Olympus SZ-PT stereomicroscope (Tokyo, Japan) and an Olympus BH2 (Tokyo, Japan), respectively. Photographs were taken with an Olympus C-35AD-4 camera (Tokyo, Japan).

Qualitative X-ray diffraction (XRD) analysis of the red colored precipitate, previously dried in an anaerobic chamber, formed by the Eerbeek sludge and by the enrichment, was performed with a Philips X'pert Diffractometer with a Cu source (Ka1 line at 0.154 nm). For these measurements, spectra were taken in the interval from 10 to 80° (2q) at a scanning speed of 2° min–1.


   RESULTS
TOP
 ABSTRACT
 INTRODUCTION
 MATERIAL AND METHODS
 RESULTS
DISCUSSION
 REFERENCES
 
Effect of Inoculum Type on Selenate Removal
Screening of different biomass types showed that anaerobic methanogenic sludge from Eerbeek had the highest selenate removal efficiency: 49% of the initial concentration (0.5 mM; 39.5 mg L–1 Se) was removed from solution in 4 d when incubated at 30°C, pH 7.5, and 20 mM lactate (Fig. 1a ). The selenate concentration (0.5 mM) leveled off after 4 d of incubation, when all lactate (20 mM) was consumed (Fig. 1b). No selenate conversion occurred in controls with killed Eerbeek sludge (Fig. 2a ).


Figure 1
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  		Fig. 1. Effect of inoculum type on (a) selenate removal and (b) lactate consumption by sludge from waste water treatment plants; (c) selenate removal and (d) lactate consumption by natural inocula. (X) Umicore, ({blacktriangleup}) Eerbeek, ({blacksquare}) Veendam, ({diamondsuit}) Nedalco, ({square}) soil mining area, ({Delta}) river sediment, ({circ}) and sea sediment.

 

Figure 2
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  		Fig. 2. Effect of electron donor (20 mM) on selenate removal by sludge Eerbeek incubated with 0.5 mM selenate. (a), (b) Selenate removal, and (c), (d) organic substrate consumption, with (•) propionate, (-) methanol, ({diamond}) acetate, ({blacksquare}) ethanol, ({blacktriangleup}) lactate, ({triangleup}) sucrose, ({square}) casein, ({circ}) glucose, (X) sterile reference and ({diamondsuit}) no substrate addition (endogenous substrate). Note that sucrose, glucose, and casein consumption was not determined.

 
Nedalco sludge gave only a 27% selenate removal in the same time interval (i.e., 100 h), whereas Umicore and Veendam sludge had even lower selenate removal efficiencies, i.e., 18 and 12%, respectively, within the same time scale and incubation conditions. Lactate was 74, 38, and 35% consumed in Nedalco, Veendam and Umicore sludge, respectively. The soil (12%) and both sediments (about 9%) had a poor selenate conversion capacity (Table 2).


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  		Table 2. Effect of inoculum source on selenate (0.5 mM) removal efficiency and rate with lactate (20 mM) as electron donor (pH 7.5; 30°C).

 
Eerbeek sludge had also the highest selenate removal rate (around 1000 µg Se g VSS–1 h–1), followed by Umicore sludge (800 µg Se g VSS–1 h–1). Comparable initial Se removal rates (about 400 to 540 µg Se g VSS–1 h–1) were obtained for Nedalco and Veendam sludge, soil, and both sediments (Table 2). Note that both the Se removal rate and Se removal efficiency were calculated for the 0- to 100-h time interval.

Effect of Electron Donor on Selenate Removal
The selenate removal efficiency by Eerbeek sludge was the highest with ethanol as the substrate (Fig. 2): ethanol-fed sludge removed 61% from the initial 0.5 mM (40 mg L–1) selenate concentration, compared to 48, 43, 36, and 36% with, respectively, lactate, glucose, propionate, and sucrose as the substrate. The maximum reduction rate was obtained for ethanol (about 1500 µg Se g VSS–1 h–1) followed by glucose, sucrose, lactate, propionate, methanol, and acetate (Table 3). Casein hydrolysate and controls without any external electron donor added did give only a low (13%) Se removal efficiency (Table 3).


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  		Table 3. Effect of electron donor on selenate removal efficiency and rate by Eerbeek anaerobic granular sludge.

 
End Product of Selenate Conversion
An orange to red coloration of the culture media, suggesting the formation of Seo, was noticed in the liquid phase of the batch tests with Eerbeek sludge supplemented with ethanol, lactate, and propionate. Initially (first 72 h of incubation), red Seo was formed in the culture media. Later (up to 100 h of incubation), red particulate spots of Seo were formed on the surface of the granules (Fig. 3a , 3b). In contrast, no color change of the culture media was noticed in the batch bottles incubated with sucrose, glucose, methanol, acetate, casein, and for the control sample.


Figure 3
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  		Fig. 3. Elemental Se (Seo) in anaerobic granular sludge incubated with 0.5 mM selenate and (a) 20 mM lactate and (b) 20 mM ethanol (30°C; pH 7.5). (c) Red Seo and biomass in Eerbeek enrichment and (d) microorganisms in Eerbeek enrichment (mixed culture). Color change and formation of red Seo in the enrichment culture supplemented with (e) 10 mM selenate and 20 mM lactate (pH 7.5, 30°C).

 
Additionally, crystalline compounds were detected by X-ray diffraction in Eerbeek biomass incubated with 20 mM lactate (Fig. 4b ). One of these minerals was calcite as previously described for Eerbeek sludge (van Hullebusch et al., 2005). Other peaks were also present in the X-ray diffraction spectra suggesting the presence of other crystalline phases. These peaks did not match with the reference spectra of achavalite (FeSe), ferroselite (FeSe2), CaSe, CuSe, CoSe, NiSe, or ZnSe.


Figure 4
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  		Fig. 4. Spectra for red elemental Se collected (a) in the enrichment of Eerbeek sludge and (b) mineral precipitates in inoculum Eerbeek sludge incubated with 20 mM lactate, and respectively, 10 and 0.5 mM selenate. Unknown peaks did not match with the reference spectra of the achavalite, ferroselite, or other metal selenides (CaSe, CuSe, CoSe, NiSe, and ZnSe).

 
Table 4 shows that based on the total Se accumulation in the granular matrix, Eerbeek sludge fed with lactate showed a higher accumulation of Se (50%) in comparison with Nedalco (22%). Low amounts of Se accumulated in Eerbeek samples incubated without any organic substrate.


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  		Table 4. Total Se accumulation (%) in Eerbeek and Nedalco anaerobic granular sludge for the initial concentration of Se of 0.5 mM.

 
Effect of Initial Selenate Concentrations
The selenate removal efficiency decreased with increasing initial selenate concentration to about 10 mM (Table 5). The highest removal efficiency of soluble Se in the batches incubated with Eerbeek sludge (Table 5; Fig. 5a ) and was achieved at an initial concentration of 0.1 mM selenate with lactate as the substrate (99%) in comparison with 0.5 mM (50%), 1 mM (36%) and 10 mM (16%). Selenate removal rates were maximal at 0.5 mM (1000 µg Se g VSS–1 h–1). Figure 6 shows that the selenate removal kinetics decreased at increasing Se concentrations.


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  		Table 5. Effect of initial Se concentration on Se removal and/or accumulation by Eerbeek sludge (20 mM lactate).

 

Figure 5
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  		Fig. 5. Effect of selenate concentration on Se removal rate ({blacksquare}) and efficiency ({diamondsuit}) with (a) intact and (b) crushed Eerbeek sludge incubated with different concentrations of selenate and 20 mM lactate.

 

Figure 6
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  		Fig. 6. Selenate removal from culture media in samples incubated with Eerbeek sludge and (a) 0.1 mM and (b) 0.5 mM selenate. ({diamondsuit}) sterile/inactive biomass and 20 mM lactate, ({blacksquare}) active biomass and 20 mM lactate, ({blacktriangleup}) and active biomass and no substrate addition (endogenous substrate).

 
Table 5 gives also the amount of Se (%) accumulated in the granules at different initial selenate concentrations. Although the Se mass balance with Eerbeek sludge was not always completely closed, most of the Se removed from solution was found back in the solid phase (Table 6).


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  		Table 6. Selenium recovery in samples incubated with Eerbeek anaerobic granular sludge.

 
The formation of selenite (SeO32–) as an intermediate was noticed after 24 h of incubation in most of the runs with elevated selenate concentrations (1 mM, 10 mM, 20 mM). For example, in Eerbeek samples incubated with 1mM selenate and 20 mM lactate, selenite (below 0.2 mM) accumulated in solution (Fig. 7 ). During its accumulation, no total soluble Se was removed from the culture media. A second feed with lactate (after 500 h) induced a further removal of both selenate and selenite.


Figure 7
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  		Fig. 7. Removal of selenate by Eerbeek sludge incubated with 1 mM selenate and 20 mM lactate. ({diamondsuit}) Selenate, (•) selenite, and ({square}) total Se.

 
Effect of Crushing Granular Sludge
The Se removal rates of crushed Eerbeek sludge incubated in the same conditions (20 mM lactate; 30°C; pH 7.5) decreased by almost 40% compared with intact granular sludge (Fig. 5a, 5b). For example, the removal rate of 0.1 mM selenate by crushed sludge occurred at a rate of 488 µg Se g VSS–1 h–1, compared to 769 µg Se g VSS–1 h–1 for intact granular Eerbeek sludge. This corresponded to selenate removal efficiencies of 32% and 99%, respectively, for crushed and intact Eerbeek sludge.

Enrichment Culture
An enrichment culture of Eerbeek granular sludge was set up with basal mineral medium supplemented with 20 mM lactate and 10 mM selenate at 30°C in the dark. Cocci-shaped cells (Fig. 3c; Fig. 3d) accumulating red elemental Se (Fig. 3c) were dominant in the enrichment culture. The enrichment reduced the initial Se concentrations of 10 mM to about 1 mM during 21 d with an efficiency of ~87% and a specific conversion rate (QSe) of 1506 µg Se(VI) g cells–1 h–1 (Fig. 8 ). The Eerbeek enrichment had a specific growth rate (µ) of 0.0125 g cells h–1 instead of a yield of (YSe) 0.083 g cells mg Se–1 with lactate as the electron donor. Growth of the enrichment culture with ethanol as the substrate was not successful (Fig. 8).


Figure 8
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  		Fig. 8. Selenate (10 mM) reduction by Eerbeek enrichment with (•) 20 mM lactate or ({diamondsuit}) 20 mM ethanol as the electron donor. Closed symbols = soluble Se concentration; open symbols = optical density.

 
During the incubation, the color of the enrichment progressively changed from transparent to orange, and finally to red in the latest phase (Fig. 3e). The formation of Seo started around the fourth day, which coincided with the exponential growth phase of the selenate-reducing bacteria. The red precipitate became more abundant during the 3 wk incubation. X-ray diffraction confirmed that the red precipitate contains Seo (Fig. 4a).


   DISCUSSION
TOP
 ABSTRACT
 INTRODUCTION
 MATERIAL AND METHODS
 RESULTS
 DISCUSSION
REFERENCES
 
Selenate Removal in Different Environments
This paper showed that selenate conversion is omnipresent in anaerobic granular sludges (Fig. 1a) and anoxic natural environments such as soil, as well as river and sea sediment (Fig. 1c). Table 2 shows that the selenate removal rates of the Eerbeek and Nedalco sludge were higher than those of the suspended biomass types, e.g., Umicore and Veendam sludge, soil, and sediment. The removal rates of the Eerbeek sludge were even higher (20%) than those of suspended sludge Umicore, exposed to high (5000 µg L–1) Se concentrations for a prolonged period of time (Picavet et al., 2003). Thus, methanogenic anaerobic granular sludge has a higher potential for soluble Se removal compared to suspended sludges and natural inocula.

The high performance of the Eerbeek sludge (Table 2) might be attributed to its granular structure, which functions as a protective barrier against metal toxicity (Lewis, 2001). Indeed, a ~40% lower selenate removal efficiency was observed when the intact structure was disrupted by crushing (Fig. 5a, 5b). Recently, Harrison et al. (2005) showed that biofilms of Pseudomonas aeruginosa or Escherichia coli are highly tolerant to a short exposure of selenite in comparison with planktonic cells. This was explained as a time-dependent tolerance mechanism developed within the biofilm, rather than an increased resistance of biofilm-grown cells.

Selenate Removal Mechanism
Adsorption and Precipitation of Selenate in Anaerobic Granular Sludge
Physicochemical processes, such as sorption and precipitation of selenate and selenite on FeS2 (Bruggeman et al., 2005), which is ubiquitously present in anaerobic sludges like Eerbeek (Gonzalez-Gil et al., 2001; van Hullebusch et al., 2005), might compete with microbially mediated reductive mechanisms (Stolz and Oremland, 1999). A solely abiotic sorption of soluble Se did, however, not occur as tests with inactive (killed) granular biomass showed no significant changes in the initial Se concentration during incubation (Fig. 2a, 6a, and 6b). Moreover, selenate is less susceptible to abiotic sorption in anaerobic conditions (Séby et al., 1998; Kuan et al., 1998) compared to selenite (de Llano et al., 1996; Bruggeman et al., 2002; Bruggeman et al., 2005). Thus, sorption is not the main mechanism responsible for the selenate removal in FeS2-rich anaerobic granular sludges.

The involvement of the sorption and reductive precipitation mechanisms in the Se removal process can, however, not be excluded as they are influenced by microbial metabolic activity. Adsorption can be a microbially mediated mechanism (White et al., 1995), offering nucleation sites for the formation of stable metal minerals (Gadd, 2004). Hockin and Gadd (2003) described the dual precipitation of Se and S by sulfate-reductive bacteria biofilms as a biologically mediated pathway. Microbial metabolism can also create low redox conditions (White et al., 1998), which may in turn catalyze the adsorption (Bruggeman et al., 2005) and subsequently, reductive precipitation of metals and metalloids. If so, it is expected that under strongly reducing conditions, Se is precipitated as a selenide (e.g., FeSe or FeSe2) (Herbel et al., 2003; Hockin and Gadd, 2006) or as metal sulfide (e.g., FeSeS) (Séby et al., 1998; Belzile et al., 2000). Although X-ray diffraction analysis confirmed indeed the presence of Se-containing mineral precipitates in the Eerbeek granules (Fig. 4b), it could not be identified as one of the iron selenides compounds (e.g., FeSe, FeSe2). Comparing the solubility products (pKsp) of iron sulfide and iron selenide (pKsp ~18 and 26 for FeS and FeSe, respectively) suggests that Se could be incorporated into sulfide minerals such as FeS and the formation of FeSeS is thermodynamically favorable (Belzile et al., 2000). More investigations are required to establish the exact contribution of both physicochemical and microbial mechanisms involved in selenate removal from solution in FeS2 rich anaerobic granular sludges.

Microbial Dissimilatory Reduction
The kinetics of selenate removal with different electron donors (Fig. 2), the consumption of organic substrates (Table 2; Fig. 1b; Fig. 2c, 2d), and the formation of Seo (Fig. 4a) and selenite (Fig. 7) strongly suggests the occurrence of a microbial selenate reduction mechanism in the anaerobic granular sludge Eerbeek. It has been shown that soluble Se (VI) can be reduced via selenite (SeO32–) or directly to insoluble Seo (Macy et al., 1993; Stolz and Oremland, 1999) and further to Se2– (Herbel et al., 2003). Selenite (SeO32–) was an intermediary product in the selenate (SeO42–) reduction tests performed with Eerbeek granular sludge (Fig. 7), but not with the microbial enrichment (Fig. 8). Due to its transitory accumulation, the formation of selenite is not the rate-limiting step of the microbial reductive Se transformation.

Organic substrates are electron donors in dissimilatory reduction processes (Lovley, 1993; Stolz and Oremland, 1999) and the selection of a suitable one is important for achieving an effective selenate removal. Figure 2a shows that ethanol is an effective substrate for selenate removal from solution and its accumulation in the solid phase (Table 4). This contrasts work with Se-contaminated sediments, where ethanol was found to be inefficient for removing Se (Nobbs et al., 1997; Caroll, 1999). Formation of insoluble Seo (Fig. 3; Fig. 4a) has profound practical implications, as Seo is a nontoxic product with commercial value. This opens perspectives to design a continuous bioreactor, treating selenate-contaminated wastewater ranging from 0.1 mM to 10 mM at pH 7.5 and 30°C, using methanogenic granular sludge as inoculum and ethanol as electron donor. The conversion of selenate into biologically nonavailable forms (i.e., Seo) (Fig. 4a) effectively decreases the Se concentrations in contaminated waters (Wang and Chen, 2003; Wu, 2004). However, it also raises further practical issues concerning the chemical stability of the formed Seo (Linkson, 1992; Simonton et al., 2000), its separation from the liquid phase (Fig. 3d), and the extraction of accumulated Se from the solid phase (i.e., granules). Recently, it has been shown that Seo can be easily removed from the aqueous phase by chemical precipitation with FeCl3, followed by centrifugation and ultrafiltration of the precipitate (Kashiva et al., 2000). For the Se formed in granular sludge (Table 5), different technologies will need to be applied; chemical or bio-leaching as done with heavy metals (Rulkens et al., 1989) could be employed. More research is needed to develop the most cost-effective Seo solid separation process.

Characterization of Selenate-Reducing Enrichment
The direct involvement of microorganisms in the selenate removal process was further evidenced by the enrichment of a microbial consortium from the anaerobic granular sludge Eerbeek (Fig. 3d; Fig. 8). Few kinetic studies on selenate-reducing bacteria have been performed to date. Promising metabolic performances for selenate reduction were found for Bacillus sp. SF-1 (Fujita et al., 2002), mixed cultures isolated from industrially Se-contaminated sediments (Nobbs et al., 1997; Caroll, 1999), and Halomonas Strain MDP-51, isolated from a Se-contaminated hypersaline evaporation pond (de Souza et al., 2001). The growth rate (µ) was almost similar, but yields (YSe) of the enrichment from the Eerbeek sludge was 5 times higher than that obtained by Caroll (1999) in his work on Se-rich contaminated sediments. Particularly, the specific conversion rate (QSe) of the enrichment from methanogenic granular sludge was almost two times higher than that of the mixed culture isolated from Se-contaminated sediments (Table 7). The variability of the analytical procedures (e.g., biomass determination) and the differences of the experimental setup (batch or bioreactor) or source of biomass (pure or mixed cultures) hamper the comparison of the present data with those from previous studies.


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  		Table 7. Metabolic parameters of different inocula during selenate removal from solution.

 
So far, dissimilatory reduction of soluble Se species to Seo has been reported to be performed mostly by bacteria isolated from natural environments: Thauera selenatis (Macy et al., 1993), Desulfovibrio desulfuricans (Tucker et al., 1998), Desulfomicrobium norvegicum (Hockin and Gadd, 2003), Sulfurospirillum barnesii (Herbel et al., 2003), or Bacillus selenitoreducens (Blum et al., 1998). The microorganisms isolated from the anaerobic granular sludge Eerbeek have not been yet identified, but seem to possess the necessary metabolic abilities for reducing high concentrations of Se (10 mM) to Seo (Fig. 4a; Fig. 8). Moreover, the microbial enrichment from Eerbeek sludge kept the selenate removal rate at a high value in comparison with granular inocula, and clearly offered the advantage of almost 90% conversion and a nontoxic end product (i.e., Seo), which, subsequently, can be separated from the liquid phase (Kashiva et al., 2000). The identification of the bacterial strains involved in selenate reduction in granular anaerobic sludge as well as the establishment of their metabolic performance remains a challenge for future biotechnological applications.


   ACKNOWLEDGMENTS
 
This research was supported by the IHP Programme of the European Commission, via the Marie Curie Training Site "Heavy Metals and Sulfur" (MCFH-1999-00950). The authors are grateful to Marcel Giesbers, who performed the XRD analysis.


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