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Application of a mer-lux biosensor for estimating bioavailable mercury in soil
Soil Biology & Biochemistry 32 (2000) 639±646
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Application of a mer-lux biosensor for estimating bioavailable mercury in soil Lasse D. Rasmussen a, Sùren J. Sùrensen a,*, Ralph R. Turner b, Tamar Barkay c a
Department of General Microbiology, University of Copenhagen, Sùlvgade 83H, DK-1307 Copenhagen K, Denmark b Frontier Geosciences, Seattle, WA 98106, USA c Deparment of Biochemistry and Microbiology, Cook College, Rutgers University, New Brunswick, NJ 08901-8525, USA Accepted 1 October 1999
Abstract A previously described bioassay using a mer-lux gene fusion for detection of bioavailable mercury was applied for the estimation of the bioavailable fraction of mercury in soil. The bioavailable fraction is de®ned here as being part of the water leachable fraction. Due to masking of light emission of soil particles leachates had to be cleaned prior to assays. Filtration of leachates through nitro-cellulose ®lters using pressure resulted in an underestimation of bioavailable mercury. Gravity ®ltration and centrifugation showed elevated (as compared with untreated leachate) and very similar responses. The utility of the mer-lux biosensor assay was tested by relating measurements of bioavailable and total mercury to the response of the soil microbial community to mercury exposure. Two dierent soil types (an agricultural and a beech forest soil) were spiked with 2.5 mg Hg(II) gÿ1 in microcosms and the frequency of mercury resistant heterotrophs and changes in community diversity, de®ned as the number of dierent 16S rDNA bands observed in DGGE gels, were monitored. In the agricultural soil the initial concentration of bioavailable mercury was estimated to be 40 ng gÿ1. This concentration did not change during the ®rst 3 d and coincided with increased degrees of resistance and a decrease in diversity. The concentration of bioavailable mercury decreased subsequently rapidly and remained just above the detection level (0.2 ng gÿ1) for the remainder of the experiment. As a possible consequence of the decreased selection pressure of mercury, the resistance and diversity gradually returned to pre-exposure amounts. In the beech forest soil the concentration of bioavailable mercury was found to be about 20 ng gÿ1 throughout the experiment. This concentration did not at any time result in changes in resistance or diversity. This study showed that the bioassay using the mer-lux biosensor is a useful and sensitive tool for estimation of bioavailable mercury in soil. 7 2000 Elsevier Science Ltd. All rights reserved.
1. Introduction The most commonly used method for estimation of environmental risk due to heavy metal pollution is quanti®cation of total metals after digestion by strong acids and chemical analysis. However, this method gives little idea of the bioavailability of metals and their potential toxicity. Several investigators have attempted to measure the bioavailability of heavy
* Corresponding author. Tel.: +45-3532-2053; fax: +45-35322040. E-mail address: [email protected] (S.J. Sùrensen).
metals. Often the bioavailable fraction is de®ned as being the solvent extractable fraction in the total concentration, e.g. weak acid (0.5 N HCl) extraction (Stone and Marsalek, 1996). Sterckerman et al. (1996) compared the concentration of metals extracted from soil using dierent solvents with concentrations in plants growing in the contaminated soils. They found close correlation between the water-extractable fraction and those taken up by plants. Accumulation in zooplankton and ®sh muscles has been used as indirect indicators of mercury bioavailability (Slotton et al., 1995). Others used the total concentration of methylated mercury as a measure of bioavailability (Regnell and Tunlid, 1991). None of the methods mentioned
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above can account for the fraction of mercury that is available for microorganisms. The issue of bioavailability is very critical with respect to mercury contamination. Bioavailable mercury, Hg(II), is the substrate for the formation of the extremely toxic methyl-mercury, the most toxic mercury form that is accumulated and concentrated in the foodchain (Barkay et al., 1992). In the environment the rate of methylation is dependent on the concentration of Hg(II) (Compeau and Bartha, 1985; Gilmour et al., 1992). At present no analytical method can distinguish bioavailable forms of mercury. Biosensors consisting of bacteria that contain gene fusions between the regulatory region of the mer operon (merR ) and bacterial luminescence genes (luxCDABE ) quantitatively respond to Hg(II) (Selifonova et al., 1993; Barkay et al., 1997). The mer promotor is activated when Hg(II), present in the cytoplasm of the biosensor bacterium, binds to MerR resulting in transcription of the lux genes and subsequent light emission. This response is quantitative. At higher concentration of Hg(II) mer promotors are activated to a greater extent resulting in higher quantities of light emission. Rasmussen et al. (1997) improved the sensitivity of this bioassay to detect approximately 1.4 ng Hg(II) lÿ1 and the method has been applied for measurements of bioavailable mercury in natural waters (Barkay et al., 1998; Turner et al., unpubl.). Our aim was to develop a method for the application of the bioassay to the measurement of bioavailable mercury in soil. The utility of this method was tested by relating measurements of bioavailable and total mercury to the response of the soil microbial community to mercury exposure in two dierent soils, an agricultural soil and a beech forest soil. 2. Materials and methods 2.1. Bacterial strains, plasmids, growth and cell preparation The strains used were two mer-lux derivatives of Escherichia coli HMS174, one containing plasmid pRB28 (Selifonova et al., 1993) and the other with a constitutive mutant of pRB28, pRB27 (Barkay et al., 1997). The constitutive mutant was used in all assays as a control to assure that light emission was not modulated by assay conditions (Barkay et al., 1997). Cultures were grown in LB medium using Kanamycin (Km) (50 mg mlÿ1) for selection of plasmids. Growth and preparation of cells for mer-lux assays were as described by Selifonova et al. (1993). The optical density of cell suspensions in 67 mM phosphate buer (pH 6.8) was adjusted to A660 corresponding to approximately 2 108 cells mlÿ1.
2.2. Soil samples and microcosm design The soil used for the development of the soil merlux assay was collected from a garden farm in Kingston, TN, USA. Two dierent soil types were used in microcosm experiments, an agricultural soil (no pesticides or fertiliser have been used for at least 20 yr) with crop change every year collected near Roskilde, Denmark and a beech forest soil from Grib Skov, Denmark (Table 1). All soils were sieved (mesh size 2 mm) and air dried at room temperature over night. Water was added to 10% (v/w) of dry weight. Mercury as HgCl2 was added to the soils with the water, water alone was added to control samples. Microcosm soils and the garden soil were spiked with 2.5 and 1 mg Hg(II) gÿ1 soil, respectively. After addition of mercury and water the soils were placed in ziplock bags and mixed thoroughly by applying manual pressure to the outside of the bag. Samples were left at room temperature for 45 min prior to leaching. Three microcosms for each treatment consisting of 50 g soil, placed in 100 ml glass beakers in ziplock bags to minimise water evaporation. Microcosms were incubated at 248C. The entire microcosm was transferred to a ziplock bag prior to each sampling, samples were obtained as described above and the remaining content of the bag returned to beakers. All glassware used were acid rinsed using 2N HNO3 and several volumes of distilled water. Samples were collected from each of the three parallel microcosms at every sampling time. 2.3. Preparation of soil leachate At least 1 g of soil (wet weight) was mixed with 10 volumes (w/v) of sterile double distilled water (ddH2O) in a 300 ml Erlenmeyer ¯ask and the ¯ask was shaken horizontally at 300 rpm at room temperature. Experiments to optimise the period of shaking showed decreased amounts of mercury in leachates when shaking exceeded 15 min (data not shown), and this shaking period was therefore chosen for all further experiments. Large soil particles were removed from leachates prior to assays by ®ltration or centrifugation. Table 1 Soil characteristics
Total C % Total N % C/N ratio Ammonium mg N/g dry soil Nitrate mg N/g dry soil Water holding capacity % pH
Agricultural
Beech Forest
1.3 0.19 6.86 0.08 7.42 24 6.6
1.7 0.17 9.76 1.26 6.64 47 3.8
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Filtrations were performed using acid rinsed millipore ®ltration setups. All ®lters used were nitro-cellulose ®lters (25 or 47 mm dia). Filtration through various poresizes, 0.22, 0.45, 0.8, 1.2, 3.0 and 8 mm, was attempted. In another approach soil leachates were ®ltered through Watman ®lter paper (No. 40) placed in glass funnels. Filtrates were collected in 300 ml Erlenmeyer ¯asks. Soil leachates were centrifuged at 12,000g for 10 min at 48C. Immediately after centrifugation or ®ltration the cleaned leachate was transferred to acid rinsed scintillation vials, diluted with sterile ddH2O, and immediately transferred to the assay tubes. In microcosm experiments centrifugation was used for the preparation of leachate. 2.4. Mer-lux assays in soil leachates A concentrated mix (Barkay et al., 1998) of assay constituants (180 ml) was added to soil leachates immediately prior to assays. Assays were performed in 20 ml glass scintillation vials for detection of light by scintillation counting or in 6 ml polystyrene tubes (Falcon, NJ, USA) for luminometer counting. The ®nal assay medium consisted of pyruvate (5 mM); Na,Kphosphate buer (67 mM PO4; 34 mM Na; 33 mM K; pH 6.8) and (NH4)2SO4 (91 mM). For quanti®cation of bioavailable mercury, 1.72 ml of appropriate dilutions of soil leachate in water were added to a ®nal volume of 1,9 ml and assays were initiated by addition of 0.1 ml biosensor cell suspension (®nal concentration of 107 cells mlÿ1). Light emission was recorded as either counts per min (cpm) in the single photon count mode of a Tri-Carp 2500 TR (Packard Instruments, Meriden, CT) scintillation counter (counter setup: count time per sample 0.5 min, 20±30 cycles, no background correction, SPC %HV: 60); or Relative Light Units (RLU) per 30 s using a BG-P Portable luminometer (MGM instruments, Hamden, USA). Luminescence measurements were taken every 5±10 min for a period of 70±90 min. 2.5. Estimation of bioavailable and total mercury The mer-lux expression factors (log quanta minÿ1) were calculated from the slopes of light emission curves as described by Barkay et al. (1998). A regression between expression factors and mercury concentration obtained from assays performed in ddH2O containing known concentrations of Hg(II) was used to calculate bioavailable mercury concentrations in soil leachates. Assays employed 107 cells of the biosensor mlÿ1 to give a linear response between Hg(II) concentration and expression factors in the concentration range of 0.3±1 nM (Rasmussen et al., 1997). Leachates
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were diluted to give expression factors that fell within this concentration range. Total mercury in soil microcosms was measured using a Jerome 431-x Mercury Vapor Analyser (Arizona Instruments, Phoenix, USA) using soil method 2 as described by Kriger and Turner (1995). 2.6. Enumeration of CFU One g of soil was added to a test tube containing 9 ml 1% NaCl in dist. water and this suspension was vortexed at maximum velocity for 60 s. Appropriate 10-fold dilutions (0.1 ml) were plated on LB agar plates containing the fungicide Natamycin 25 mg mlÿ1 (Merck) (Pedersen, 1992). Mercury-resistant heterotrophs were enumerated on similar medium prepared with 10 mg Hg(II) as HgCl2 mlÿ1. All plates were incubated at 24 8C for 4 d prior to enumeration. 2.7. Bacterial diversity Diversity analysis of the microcosm bacterial community were performed at every sampling point by extracting total DNA, PCR ampli®cation of 16S rDNA fragments followed by sequence separation by denaturing gradient gel electrophoresis (DGGE) (Muyzer et al., 1993). DNA extractions were carried out as described by Porteous et al. (1994), except that the sonication time was reduced from 3 min to 10 s since tests showed that DNA was rapidly lost with longer sonication time (data not shown). Removal of humic acids was performed by gel electrophoresis (0.7% low melting SeaPlaque agarose) for 1 h at 125 V. Following electrophoresis gel blocks containing the DNA were cut from the gel and stored at ÿ208C in Eppendor tubes. Immediately prior to PCR ampli®cation the gel blocks were melted at 688C for 5 min, 3 volumes of ddH2O were added and samples were incubated for 20 min at 688C. PCR was performed with `ready to go' PCR beads as described by the manufacturer (Pharmacia): 1 ml of DNA sample was mixed with 675 nl of each primer (for primer sequence see Muyzer et al. (1993)) and sterile ®ltered ddH2O to a total volume of 25 ml. Ampli®cation was achieved by one cycle of: 948C 4 min, 608C 1 min, 728C 1 min followed by 34 cycles of: 948C 1 min, 608C 1 min, 728C 1 min and the last cycle was followed by 8 min at 728C. Dierent 16S rDNA sequences were separated using DGGE as described by Muyzer et al. (1993). The D GENE System equipment (BIO-RAD) was used for the preparation of gels and electrophoresis. Gels were stained with SYBR Green (1:10,000 dilution, Molecular Probes, Eugene, USA) for 1 h. The number of bands in DGGE gels, counted manually from Polaroid pictures were used as measure of bacterial diversity.
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3. Results 3.1. Optimization of the mer-lux assay for the analysis of bioavailable mercury in soil leachate In order to avoid blocking of light and changes in the concentration of bioavailable mercury during the assay, soil particles were removed from the leachates by ®ltration or centrifugation (see below). To evaluate the eect of ®ltration on the concentration of bioavailable mercury, leachates of the garden soil (supplemented with 1 mg Hg(II) gÿ1 soil prior to leaching) were ®ltered through ®lters with dierent pore sizes. The biosensor response declined with decreasing pore size (Fig. 1A) resulting in total inhibition of mer-lux induction with ®ltration through a pore size smaller than 3.0 mm. The response in all ®ltrates was lower than that of the raw un®ltered leachate. Thus, it seems that ®ltration sequestered
bioavailable mercury. This could be due to loss of mercury by binding to ®lters or soil particles collected on the ®lter or to the release of material that binds mercury during ®ltration. That the latter was the case was shown by comparing induction in dist. H2O and the 0.22 mm ®ltrate to which 0.75 nM Hg(II) (as HgNO3) were added. The much lower response of the leachate (Fig. 1B) indicates that ®ltration under pressure released substances that reduced mercury bioavailability. Cell lysis that might have occurred during ®ltration is the likely source of these substances. This is supported by the fact that increasing pore size resulted in less inhibition of mer-lux responses (Fig. 1A) since the larger the pores the lower is the pressure built-up during ®ltration and the more likely are cells to pass intact through the ®lter. Indeed, when pressure was varied during ®ltration induction of merlux was totally abolished when high pressure was applied while low pressure resulted in a signi®cant re-
Fig. 1. The eect of ®ltration of soil leachate on the bioavailability of mercury. (A) Induction of mer-lux in soil (spiked with 1 mg Hg gÿ1 prior to leaching). Leachates were ®ltered through ®lters with increasing pore size. (B) Induction of mer-lux in distilled water and a ®ltrate (0.22 mm) of soil leachate spiked with 15 ng Hg(II) mlÿ1.
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sponse that was nevertheless decreased relative to that of un®ltered leachates (data not shown). These results suggested that ®ltration by forcing leachates through nitro-cellulose ®lters might result in an underestimation of bioavailable mercury concentrations. The response of assays in leachate obtained by gravity ®ltrating through Watman ®lter paper (No. 40) was compared with leachate cleaned by centrifugation. Both treatments showed elevated and very similar responses (Fig. 2A). Expression factors indicating the rate of increase in light emission (Barkay et al., 1998) were calculated to be 0.119 and 0.112 for gravity ®ltration and centrifugation respectively. While that of the untreated leachate was only 0.081. These results con®rm that soil particles had to be removed to prevent underestimation of the amount of bioavailable mercury in soil leachates. Assays using the constitutive mer-lux derivative HMS174/pRB27, revealed that underestimation was due to masking of light in raw leachate (Fig. 2B).
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Gravity ®ltration was slow (approximate ¯ow rate 10 ml leachate in 60 min), making this procedure extremely time consuming. Centrifugation was therefore used for soil leachate preparation in all following experiments. Furthermore, since centrifugation also removed indigenous bacteria from the leachate and previous work (Rasmussen et al., 1997) showed that bacterial density in the assay medium aect the assay's sensitivity, this method was favored over gravity ®ltration. The constitutive mutant derivative of the biosensor, strain E. coli HMS174/pRB27 (Barkay et al., 1997) showed that light emission was not quenched in any of the soil leachates employed here (data not shown). 3.2. Microcosm experiments 3.2.1. Bioavailable and total mercury The utility of the mer-lux biosensor was tested by relating measurements of bioavailable and total mer-
Fig. 2. Comparison of biosensor response in gravity ®ltered, centrifuged and raw untreated soil leachates. Numbers in parenthesis are calculated expression factors (see text for further information).
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cury to the response of the soil microbial community to mercury exposure in soil microcosms using two distinctly dierent soils. Concentrations of bioavailable mercury were measured over a period of 15 d after Hg addition. Both soils demonstrated a decline in bioavailable mercury, albeit with dierent patterns (Fig. 3A). The initial decrease that is apparent between d 0 and d 1 after spiking might have been attributed to an overesti-
mation of bioavailable mercury on d 0 due to the fact that an equilibrium between mercury and soil binding sites may not yet have been established after 45 min. After this initial decrease the agricultural soil retained a constant concentration of bioavailable mercury of about 40 ng gÿ1 soil, followed by a considerable decrease to 0.3 ng gÿ1 between d 3 and d 5. The concentration of bioavailable mercury stayed just above the detection limit (0.2 ng gÿ1) throughout the rest of
Fig. 3. The relationships of mercury bioavailability to the response of the soil microbial community. Microcosm experiments using an agricultural (q) or a beech forest soil (*) were set up to relate bioavailable mercury (A) to the development of bacterial resistance to mercury measured as percent Hg resistant bacteria of all culturable bacteria (B) and to bacterial diversity measured as number of 16S rDNA bands counted on DGGE gels (C).
L.D. Rasmussen et al. / Soil Biology & Biochemistry 32 (2000) 639±646
the experiment. This decrease in bioavailable mercury concentration was not observed in the beech forest soil in which about 20 ng Hg gÿ1 soil was found for the duration of the experiment (15 d; Fig. 3A). In both soils only a very small fraction of the total added mercury (2.5 mg gÿ1) was bioavailable and the transition from available to unavailable forms was rapid with only 2% available Hg an hour after addition of mercury to the soil (Fig. 3A). The concentration of total mercury did not change in the soils during the experiment. In the agricultural soil total mercury was 2.7 20.2 mg Hg gÿ1 dry soil on d 0 and 2.6 2 0.1 mg Hg gÿ1 at the end of the experiment on d 15. During the same period, beech forest soil had 3.220.2 and 3.320.5 mg Hg gÿ1 dry soil, respectively. The background concentration of total mercury was similar in both soils (0.19 and 0.27 mg gÿ1 in agricultural and beech forest soil, respectively), the cause of the dierent concentrations of mercury after addition is not known. 3.2.2. Frequency of mercury resistant bacteria In the agricultural soil the amount of mercury resistance began to increase after the ®rst day of exposure. The frequency of mercury-resistant isolates increased rapidly reaching its peak on d 4 at 7.122% (Fig. 3B). A gradual decline followed until a plateau was reached between d 7 and d 11, where the % Hg resistance frequency remained, slightly above pre exposure quantities, for the duration of the experiment. No eect of the added mercury was observed on the frequency of mercury resistant bacteria in the beech forest soil with a constant fraction of approx. 0.004% of the total CFU growing on the mercury supplemented medium (Fig. 3B). 3.2.3. Bacterial diversity Diversity was evaluated by the number of bands visualised in DGGE gels after electrophoresis of PCR ampli®cation product obtained using primers speci®c to all eubacterial 16S rRNA genes. This analysis provides information on the diversity of the eubacterial community at large (i.e. without the need to culture soil bacteria) (Muyzer et al., 1993; Heuer and Smalla, 1997). Diversity of the indigenous bacteria in the two soils responded dierently to mercury exposure (Fig. 3C). In the agricultural soil diversity decreased rapidly in the ®rst 4 d of the experiment with the average number of 16S rDNA bands declining from 26 to 20. After this minimum the number of bands gradually increased for the rest of the experiment reaching the initial number on d 15. In the beech forest soil no eect of mercury exposure was observed on the number of 16S rDNA bands, i.e. diversity, which were about 27 for the duration of the experiment.
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4. Discussion This investigation showed that the mer-lux bioassay is a useful tool for quanti®cation of bioavailable mercury in soil leachate obtained from three dierent soils. In this study the water-leachable fraction of mercury is referred to as being potentially bioavailable. This is in agreement with Sterckeman et al. (1996) who found a strong correlation between water-extractable mercury and accumulation of mercury in plants. Bacterial activity in the soil is located in niches characterized by the availability of water, and the fact that Hg(II) needs to be in aqueous solution in order to be transported to the cytoplasm is supporting the assumption that bioavailable mercury in soil is water leachable. Results of experiments designed to optimize the merlux bioassay in soil leachate showed the biosensor response might be biased by the presence of soil particles. Forced ®ltration to remove soil particles, resulted in an underestimation of bioavailable mercury (Fig. 1A). Assays performed in water and ®ltrate spiked with 0.75 nM Hg(II) showed that this was due to mercury-binding ligands that were released during ®ltration, probably by pressure-induced cell lysis (Fig. 1B). This corresponds well with earlier ®ndings that dissolved organic carbon (DOC) considerably decreases biosensor response (Barkay et al., 1997). That removal of soil particles from leachates was needed to avoid underestimation of the concentration of bioavailable mercury was shown by comparing biosensor responses in assays performed on gravity ®ltered and centrifuged leachates with assays in raw untreated leachate (Fig. 2). Assays performed with the constitutive mutant showed that this was most probably due to shading by soil particles (Fig. 2B). Little is known about the toxicity of mercury associated with colloidal and ®ne particles. Since these fractions of the total mercury will be eliminated from the soil leachate by centrifugation eventual bioavailability of these will not be detected by this assay. The microcosm experiment showed that the microbial response to mercury observed as development of mercury-resistant bacteria and lowering of diversity was correlated to changes in concentrations of bioavailable mercury (Fig. 3). Both these responses indicate toxicity and are well documented eects of heavy metal contaminations in soil (Roane and Kellogg, 1996; Ranjard et al., 1997; Smit et al., 1998). The heavy metal concentration at which a bacterial response is elicited may vary greatly depending on the metal and the soil type (BaÊaÊth, 1989). In our study, the dierences in response of the two bacterial communities in the two soil types to dosing with equal amounts of total mercury, con®rms that bioavailable rather than total mercury is the factor controlling microbial responses. This is in good agreement with a study by Ran-
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jard et al. (1997) who after spiking four dierent soil types with the same mercury concentration found that after incubation the proportion of resistant bacteria varied from 0.4 to 36% . Many abiotic factors may aect the bioavailability of heavy metals in soil e.g. clay content, pH, dissolved organic carbon, root exudates (BaÊaÊth, 1989; Barkay et al., 1997; Giller et al., 1998). The main dierences in the soil variables (Table 1) between the agricultural and the beech forest soil is pH and ammonium (NH4) content. Soil pH is often found to have a large in¯uence on metal availability due to its strong eect on metal solubility and speciation. Thus, a decrease in pH usually results in increased availability (Giller et al., 1998). This seems not to be the case in this study where the pH in the beech forest soil was almost 3 units lower than that in agricultural soil (Table 1). The high ammonium content in the beech forest soil (Table 1) may be an important factor decreasing bioavailability of mercury. Experiments with varying the concentration of the dierent bioassay medium constituents have shown that the sensitivity of the assay is decreased by increasing the ammonia concentration (unpublished data). The results from beech forest soil shows that although bioavailable mercury was present in detectable concentrations in the soil, there was no enrichment of resistant bacteria or decline in community diversity (Fig. 3). These results suggest that bioavailable mercury at the detected concentration may not have been toxic to indigenous bacteria in beech forest soil. The bioassay requires the presence of bioavailable Hg(II) in the biosensor cytoplasm. This species of mercury is the substrate for mercury methylation by soil bacteria (Beckert et al., 1974). The fact that detectable bioavailable mercury in beech forest soil was not toxic suggests that a potential for mercury methylation exists at subtoxic concentrations of mercury. Acknowledgements L.D.R. is supported by ``Centre for biological processes in contaminated soil and sediment'' (www.biopro.dk) under The Danish Environmental research programme. The authors wish to thank Pia Kringelum and Inge E. Larsen for excellent technical assistance. References Barkay, T., Turner, R., Saouter, E., Horn, J., 1992. Mercury biotransformation and their potential for bioremediation of mercury contamination. Biodegradation 3, 147±159. Barkay, T., Gillman, M., Turner, R.R., 1997. Eects of dissolved organic carbon and salinity on bioavailability of mercury. Applied and Environmental Microbiology 63, 4267±4271. Barkay, T., Turner, R.R., Rasmussen, L.D., Kelly, C.A., Rudd, J.W.M., 1998. Luminescence facilitated detection of bioavailable
mercury in natural waters. In: LaRossa, R.A. (Ed.), Bioluminescence Methods and Protocols, Methods in Molecular Biology, Vol. 102. Humana Press, Totowa. Beckert, W.F., Moghissi, A.A., Au, F.H.F., Bretthauer, E.W., McFarlane, J.C., 1974. Formation of methylmercury in a terrestrial environment. Nature 249, 674±675. BaÊaÊth, E., 1989. Eect of heavy metals in soil on microbial processes and populations: a review. Water Air and Soil Pollution 47, 335± 379. Compeau, G.C., Bartha, R., 1985. Sulfate-reducing bacteria: principal methylators of mercury in anoxic estuarine sediment. Applied and Environmental Microbiology 50, 498±502. Giller, K.E., Witter, E., McGrath, S.P., 1998. Toxicity of heavy metals to microorganisms and microbial processes in agricultural soil: a review. Soil Biology & Biochemistry 30, 1389±1414. Gilmour, C.C., Henry, E.A., Mitchell, R., 1992. Sulfate stimulation of mercury methylation in freshwater sediments. Environmental Science and Technology 26, 2281±2285. Heuer, H., Smalla, K. 1997. Application of denaturing gradient gel electrophoresis and temperature gradient gel electrophoresis for studying soil microbial communities. In: van Elsas, J.D., Trevors, J.T., Wellington, E.M.H. (Eds.), Modern Soil Microbiology. Marcel Dekker, New York. Kriger, A.A., Turner, R.R., 1995. Field analysis of mercury in water, sediment and soil using static headspace analysis. Water Air and Soil Pollution 80, 1295±1304. Muyzer, G., de Waal, E.C., Uitterlinden, A.G., 1993. Pro®ling of complex microbial populations by denaturing gradient gel electrophoresis of polymerase chain reaction-ampli®ed genes coding for 16S rRNA. Applied and Environmental Microbiology 59, 695± 700. Pedersen, J.C., 1992. Natamycin as a fungicide in agar media. Applied and Environmental Microbiology 58, 1064±1066. Porteous, L.A., Armstrong, J.L., Seidler, R.J., Watrud, L.S., 1994. An eective method to extract DNA from environmental samples for polymerase chain reaction ampli®cation and DNA ®ngerprint analysis. Current Microbiology 29, 301±307. Ranjard, L., Richaume, A., Jocteur-Monrozier, L., Nazaret, S., 1997. Response of soil bacteria to Hg(II) in relation to soil characteristics and cell location. FEMS Microbial Ecology 24, 321±331. Rasmussen, L.D., Turner, R.R., Barkay, T., 1997. Cell-densitydependent sensitivity of a mer-lux bioassay. Applied and Environmental Microbiology 63, 3291±3293. Regnell, O., Tunlid, A., 1991. Laboratory study of chemical speciation of mercury in lake sediment and water under aerobic and anaerobic conditions. Applied and Environmental Microbiology 57, 789±795. Roane, T.M., Kellogg, S.T., 1996. Characterization of bacterial communities in heavy metal contaminated soils. Canadian Journal of Microbiology 42, 593±603. Selifonova, O., Burlage, R., Barkay, T., 1993. Bioluminescent sensors for detection of bioavailable mercury(II) in the environment. Applied and Environmental Microbiology 59, 3083±3090. Slotton, D.G., Reuter, J.E., Goldman, C.R., 1995. Mercury uptake patterns of biota in a seasonally anoxic northern California reservoir. Water Air and Soil Pollution 80, 841±850. Smit, E., Wolters, A., van Elsas, J.D., 1998. Self-transmissible mercury resistance plasmids with gene-mobilizing capacity in soil bacterial populations: in¯uence of wheat roots and mercury addition. Applied and Environmental Microbiology 64, 1210±1219. Sterckeman, T., Gomez, A., Ciesielski, H., 1996. Soil and waste analysis for environmental risk assessment in France. Science of the Total Environment 178, 63±69. Stone, M., Marsalek, J., 1996. Trace metal composition and speciation in street sediment: Sault Ste. Marie, Canada. Water Air and Soil Pollution 87, 149±169.
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Application of a mer-lux biosensor for estimating bioavailable mercury in soil Lasse D. Rasmussen a, Sùren J. Sùrensen a,*, Ralph R. Turner b, Tamar Barkay c a
Department of General Microbiology, University of Copenhagen, Sùlvgade 83H, DK-1307 Copenhagen K, Denmark b Frontier Geosciences, Seattle, WA 98106, USA c Deparment of Biochemistry and Microbiology, Cook College, Rutgers University, New Brunswick, NJ 08901-8525, USA Accepted 1 October 1999
Abstract A previously described bioassay using a mer-lux gene fusion for detection of bioavailable mercury was applied for the estimation of the bioavailable fraction of mercury in soil. The bioavailable fraction is de®ned here as being part of the water leachable fraction. Due to masking of light emission of soil particles leachates had to be cleaned prior to assays. Filtration of leachates through nitro-cellulose ®lters using pressure resulted in an underestimation of bioavailable mercury. Gravity ®ltration and centrifugation showed elevated (as compared with untreated leachate) and very similar responses. The utility of the mer-lux biosensor assay was tested by relating measurements of bioavailable and total mercury to the response of the soil microbial community to mercury exposure. Two dierent soil types (an agricultural and a beech forest soil) were spiked with 2.5 mg Hg(II) gÿ1 in microcosms and the frequency of mercury resistant heterotrophs and changes in community diversity, de®ned as the number of dierent 16S rDNA bands observed in DGGE gels, were monitored. In the agricultural soil the initial concentration of bioavailable mercury was estimated to be 40 ng gÿ1. This concentration did not change during the ®rst 3 d and coincided with increased degrees of resistance and a decrease in diversity. The concentration of bioavailable mercury decreased subsequently rapidly and remained just above the detection level (0.2 ng gÿ1) for the remainder of the experiment. As a possible consequence of the decreased selection pressure of mercury, the resistance and diversity gradually returned to pre-exposure amounts. In the beech forest soil the concentration of bioavailable mercury was found to be about 20 ng gÿ1 throughout the experiment. This concentration did not at any time result in changes in resistance or diversity. This study showed that the bioassay using the mer-lux biosensor is a useful and sensitive tool for estimation of bioavailable mercury in soil. 7 2000 Elsevier Science Ltd. All rights reserved.
1. Introduction The most commonly used method for estimation of environmental risk due to heavy metal pollution is quanti®cation of total metals after digestion by strong acids and chemical analysis. However, this method gives little idea of the bioavailability of metals and their potential toxicity. Several investigators have attempted to measure the bioavailability of heavy
* Corresponding author. Tel.: +45-3532-2053; fax: +45-35322040. E-mail address: [email protected] (S.J. Sùrensen).
metals. Often the bioavailable fraction is de®ned as being the solvent extractable fraction in the total concentration, e.g. weak acid (0.5 N HCl) extraction (Stone and Marsalek, 1996). Sterckerman et al. (1996) compared the concentration of metals extracted from soil using dierent solvents with concentrations in plants growing in the contaminated soils. They found close correlation between the water-extractable fraction and those taken up by plants. Accumulation in zooplankton and ®sh muscles has been used as indirect indicators of mercury bioavailability (Slotton et al., 1995). Others used the total concentration of methylated mercury as a measure of bioavailability (Regnell and Tunlid, 1991). None of the methods mentioned
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above can account for the fraction of mercury that is available for microorganisms. The issue of bioavailability is very critical with respect to mercury contamination. Bioavailable mercury, Hg(II), is the substrate for the formation of the extremely toxic methyl-mercury, the most toxic mercury form that is accumulated and concentrated in the foodchain (Barkay et al., 1992). In the environment the rate of methylation is dependent on the concentration of Hg(II) (Compeau and Bartha, 1985; Gilmour et al., 1992). At present no analytical method can distinguish bioavailable forms of mercury. Biosensors consisting of bacteria that contain gene fusions between the regulatory region of the mer operon (merR ) and bacterial luminescence genes (luxCDABE ) quantitatively respond to Hg(II) (Selifonova et al., 1993; Barkay et al., 1997). The mer promotor is activated when Hg(II), present in the cytoplasm of the biosensor bacterium, binds to MerR resulting in transcription of the lux genes and subsequent light emission. This response is quantitative. At higher concentration of Hg(II) mer promotors are activated to a greater extent resulting in higher quantities of light emission. Rasmussen et al. (1997) improved the sensitivity of this bioassay to detect approximately 1.4 ng Hg(II) lÿ1 and the method has been applied for measurements of bioavailable mercury in natural waters (Barkay et al., 1998; Turner et al., unpubl.). Our aim was to develop a method for the application of the bioassay to the measurement of bioavailable mercury in soil. The utility of this method was tested by relating measurements of bioavailable and total mercury to the response of the soil microbial community to mercury exposure in two dierent soils, an agricultural soil and a beech forest soil. 2. Materials and methods 2.1. Bacterial strains, plasmids, growth and cell preparation The strains used were two mer-lux derivatives of Escherichia coli HMS174, one containing plasmid pRB28 (Selifonova et al., 1993) and the other with a constitutive mutant of pRB28, pRB27 (Barkay et al., 1997). The constitutive mutant was used in all assays as a control to assure that light emission was not modulated by assay conditions (Barkay et al., 1997). Cultures were grown in LB medium using Kanamycin (Km) (50 mg mlÿ1) for selection of plasmids. Growth and preparation of cells for mer-lux assays were as described by Selifonova et al. (1993). The optical density of cell suspensions in 67 mM phosphate buer (pH 6.8) was adjusted to A660 corresponding to approximately 2 108 cells mlÿ1.
2.2. Soil samples and microcosm design The soil used for the development of the soil merlux assay was collected from a garden farm in Kingston, TN, USA. Two dierent soil types were used in microcosm experiments, an agricultural soil (no pesticides or fertiliser have been used for at least 20 yr) with crop change every year collected near Roskilde, Denmark and a beech forest soil from Grib Skov, Denmark (Table 1). All soils were sieved (mesh size 2 mm) and air dried at room temperature over night. Water was added to 10% (v/w) of dry weight. Mercury as HgCl2 was added to the soils with the water, water alone was added to control samples. Microcosm soils and the garden soil were spiked with 2.5 and 1 mg Hg(II) gÿ1 soil, respectively. After addition of mercury and water the soils were placed in ziplock bags and mixed thoroughly by applying manual pressure to the outside of the bag. Samples were left at room temperature for 45 min prior to leaching. Three microcosms for each treatment consisting of 50 g soil, placed in 100 ml glass beakers in ziplock bags to minimise water evaporation. Microcosms were incubated at 248C. The entire microcosm was transferred to a ziplock bag prior to each sampling, samples were obtained as described above and the remaining content of the bag returned to beakers. All glassware used were acid rinsed using 2N HNO3 and several volumes of distilled water. Samples were collected from each of the three parallel microcosms at every sampling time. 2.3. Preparation of soil leachate At least 1 g of soil (wet weight) was mixed with 10 volumes (w/v) of sterile double distilled water (ddH2O) in a 300 ml Erlenmeyer ¯ask and the ¯ask was shaken horizontally at 300 rpm at room temperature. Experiments to optimise the period of shaking showed decreased amounts of mercury in leachates when shaking exceeded 15 min (data not shown), and this shaking period was therefore chosen for all further experiments. Large soil particles were removed from leachates prior to assays by ®ltration or centrifugation. Table 1 Soil characteristics
Total C % Total N % C/N ratio Ammonium mg N/g dry soil Nitrate mg N/g dry soil Water holding capacity % pH
Agricultural
Beech Forest
1.3 0.19 6.86 0.08 7.42 24 6.6
1.7 0.17 9.76 1.26 6.64 47 3.8
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Filtrations were performed using acid rinsed millipore ®ltration setups. All ®lters used were nitro-cellulose ®lters (25 or 47 mm dia). Filtration through various poresizes, 0.22, 0.45, 0.8, 1.2, 3.0 and 8 mm, was attempted. In another approach soil leachates were ®ltered through Watman ®lter paper (No. 40) placed in glass funnels. Filtrates were collected in 300 ml Erlenmeyer ¯asks. Soil leachates were centrifuged at 12,000g for 10 min at 48C. Immediately after centrifugation or ®ltration the cleaned leachate was transferred to acid rinsed scintillation vials, diluted with sterile ddH2O, and immediately transferred to the assay tubes. In microcosm experiments centrifugation was used for the preparation of leachate. 2.4. Mer-lux assays in soil leachates A concentrated mix (Barkay et al., 1998) of assay constituants (180 ml) was added to soil leachates immediately prior to assays. Assays were performed in 20 ml glass scintillation vials for detection of light by scintillation counting or in 6 ml polystyrene tubes (Falcon, NJ, USA) for luminometer counting. The ®nal assay medium consisted of pyruvate (5 mM); Na,Kphosphate buer (67 mM PO4; 34 mM Na; 33 mM K; pH 6.8) and (NH4)2SO4 (91 mM). For quanti®cation of bioavailable mercury, 1.72 ml of appropriate dilutions of soil leachate in water were added to a ®nal volume of 1,9 ml and assays were initiated by addition of 0.1 ml biosensor cell suspension (®nal concentration of 107 cells mlÿ1). Light emission was recorded as either counts per min (cpm) in the single photon count mode of a Tri-Carp 2500 TR (Packard Instruments, Meriden, CT) scintillation counter (counter setup: count time per sample 0.5 min, 20±30 cycles, no background correction, SPC %HV: 60); or Relative Light Units (RLU) per 30 s using a BG-P Portable luminometer (MGM instruments, Hamden, USA). Luminescence measurements were taken every 5±10 min for a period of 70±90 min. 2.5. Estimation of bioavailable and total mercury The mer-lux expression factors (log quanta minÿ1) were calculated from the slopes of light emission curves as described by Barkay et al. (1998). A regression between expression factors and mercury concentration obtained from assays performed in ddH2O containing known concentrations of Hg(II) was used to calculate bioavailable mercury concentrations in soil leachates. Assays employed 107 cells of the biosensor mlÿ1 to give a linear response between Hg(II) concentration and expression factors in the concentration range of 0.3±1 nM (Rasmussen et al., 1997). Leachates
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were diluted to give expression factors that fell within this concentration range. Total mercury in soil microcosms was measured using a Jerome 431-x Mercury Vapor Analyser (Arizona Instruments, Phoenix, USA) using soil method 2 as described by Kriger and Turner (1995). 2.6. Enumeration of CFU One g of soil was added to a test tube containing 9 ml 1% NaCl in dist. water and this suspension was vortexed at maximum velocity for 60 s. Appropriate 10-fold dilutions (0.1 ml) were plated on LB agar plates containing the fungicide Natamycin 25 mg mlÿ1 (Merck) (Pedersen, 1992). Mercury-resistant heterotrophs were enumerated on similar medium prepared with 10 mg Hg(II) as HgCl2 mlÿ1. All plates were incubated at 24 8C for 4 d prior to enumeration. 2.7. Bacterial diversity Diversity analysis of the microcosm bacterial community were performed at every sampling point by extracting total DNA, PCR ampli®cation of 16S rDNA fragments followed by sequence separation by denaturing gradient gel electrophoresis (DGGE) (Muyzer et al., 1993). DNA extractions were carried out as described by Porteous et al. (1994), except that the sonication time was reduced from 3 min to 10 s since tests showed that DNA was rapidly lost with longer sonication time (data not shown). Removal of humic acids was performed by gel electrophoresis (0.7% low melting SeaPlaque agarose) for 1 h at 125 V. Following electrophoresis gel blocks containing the DNA were cut from the gel and stored at ÿ208C in Eppendor tubes. Immediately prior to PCR ampli®cation the gel blocks were melted at 688C for 5 min, 3 volumes of ddH2O were added and samples were incubated for 20 min at 688C. PCR was performed with `ready to go' PCR beads as described by the manufacturer (Pharmacia): 1 ml of DNA sample was mixed with 675 nl of each primer (for primer sequence see Muyzer et al. (1993)) and sterile ®ltered ddH2O to a total volume of 25 ml. Ampli®cation was achieved by one cycle of: 948C 4 min, 608C 1 min, 728C 1 min followed by 34 cycles of: 948C 1 min, 608C 1 min, 728C 1 min and the last cycle was followed by 8 min at 728C. Dierent 16S rDNA sequences were separated using DGGE as described by Muyzer et al. (1993). The D GENE System equipment (BIO-RAD) was used for the preparation of gels and electrophoresis. Gels were stained with SYBR Green (1:10,000 dilution, Molecular Probes, Eugene, USA) for 1 h. The number of bands in DGGE gels, counted manually from Polaroid pictures were used as measure of bacterial diversity.
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3. Results 3.1. Optimization of the mer-lux assay for the analysis of bioavailable mercury in soil leachate In order to avoid blocking of light and changes in the concentration of bioavailable mercury during the assay, soil particles were removed from the leachates by ®ltration or centrifugation (see below). To evaluate the eect of ®ltration on the concentration of bioavailable mercury, leachates of the garden soil (supplemented with 1 mg Hg(II) gÿ1 soil prior to leaching) were ®ltered through ®lters with dierent pore sizes. The biosensor response declined with decreasing pore size (Fig. 1A) resulting in total inhibition of mer-lux induction with ®ltration through a pore size smaller than 3.0 mm. The response in all ®ltrates was lower than that of the raw un®ltered leachate. Thus, it seems that ®ltration sequestered
bioavailable mercury. This could be due to loss of mercury by binding to ®lters or soil particles collected on the ®lter or to the release of material that binds mercury during ®ltration. That the latter was the case was shown by comparing induction in dist. H2O and the 0.22 mm ®ltrate to which 0.75 nM Hg(II) (as HgNO3) were added. The much lower response of the leachate (Fig. 1B) indicates that ®ltration under pressure released substances that reduced mercury bioavailability. Cell lysis that might have occurred during ®ltration is the likely source of these substances. This is supported by the fact that increasing pore size resulted in less inhibition of mer-lux responses (Fig. 1A) since the larger the pores the lower is the pressure built-up during ®ltration and the more likely are cells to pass intact through the ®lter. Indeed, when pressure was varied during ®ltration induction of merlux was totally abolished when high pressure was applied while low pressure resulted in a signi®cant re-
Fig. 1. The eect of ®ltration of soil leachate on the bioavailability of mercury. (A) Induction of mer-lux in soil (spiked with 1 mg Hg gÿ1 prior to leaching). Leachates were ®ltered through ®lters with increasing pore size. (B) Induction of mer-lux in distilled water and a ®ltrate (0.22 mm) of soil leachate spiked with 15 ng Hg(II) mlÿ1.
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sponse that was nevertheless decreased relative to that of un®ltered leachates (data not shown). These results suggested that ®ltration by forcing leachates through nitro-cellulose ®lters might result in an underestimation of bioavailable mercury concentrations. The response of assays in leachate obtained by gravity ®ltrating through Watman ®lter paper (No. 40) was compared with leachate cleaned by centrifugation. Both treatments showed elevated and very similar responses (Fig. 2A). Expression factors indicating the rate of increase in light emission (Barkay et al., 1998) were calculated to be 0.119 and 0.112 for gravity ®ltration and centrifugation respectively. While that of the untreated leachate was only 0.081. These results con®rm that soil particles had to be removed to prevent underestimation of the amount of bioavailable mercury in soil leachates. Assays using the constitutive mer-lux derivative HMS174/pRB27, revealed that underestimation was due to masking of light in raw leachate (Fig. 2B).
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Gravity ®ltration was slow (approximate ¯ow rate 10 ml leachate in 60 min), making this procedure extremely time consuming. Centrifugation was therefore used for soil leachate preparation in all following experiments. Furthermore, since centrifugation also removed indigenous bacteria from the leachate and previous work (Rasmussen et al., 1997) showed that bacterial density in the assay medium aect the assay's sensitivity, this method was favored over gravity ®ltration. The constitutive mutant derivative of the biosensor, strain E. coli HMS174/pRB27 (Barkay et al., 1997) showed that light emission was not quenched in any of the soil leachates employed here (data not shown). 3.2. Microcosm experiments 3.2.1. Bioavailable and total mercury The utility of the mer-lux biosensor was tested by relating measurements of bioavailable and total mer-
Fig. 2. Comparison of biosensor response in gravity ®ltered, centrifuged and raw untreated soil leachates. Numbers in parenthesis are calculated expression factors (see text for further information).
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cury to the response of the soil microbial community to mercury exposure in soil microcosms using two distinctly dierent soils. Concentrations of bioavailable mercury were measured over a period of 15 d after Hg addition. Both soils demonstrated a decline in bioavailable mercury, albeit with dierent patterns (Fig. 3A). The initial decrease that is apparent between d 0 and d 1 after spiking might have been attributed to an overesti-
mation of bioavailable mercury on d 0 due to the fact that an equilibrium between mercury and soil binding sites may not yet have been established after 45 min. After this initial decrease the agricultural soil retained a constant concentration of bioavailable mercury of about 40 ng gÿ1 soil, followed by a considerable decrease to 0.3 ng gÿ1 between d 3 and d 5. The concentration of bioavailable mercury stayed just above the detection limit (0.2 ng gÿ1) throughout the rest of
Fig. 3. The relationships of mercury bioavailability to the response of the soil microbial community. Microcosm experiments using an agricultural (q) or a beech forest soil (*) were set up to relate bioavailable mercury (A) to the development of bacterial resistance to mercury measured as percent Hg resistant bacteria of all culturable bacteria (B) and to bacterial diversity measured as number of 16S rDNA bands counted on DGGE gels (C).
L.D. Rasmussen et al. / Soil Biology & Biochemistry 32 (2000) 639±646
the experiment. This decrease in bioavailable mercury concentration was not observed in the beech forest soil in which about 20 ng Hg gÿ1 soil was found for the duration of the experiment (15 d; Fig. 3A). In both soils only a very small fraction of the total added mercury (2.5 mg gÿ1) was bioavailable and the transition from available to unavailable forms was rapid with only 2% available Hg an hour after addition of mercury to the soil (Fig. 3A). The concentration of total mercury did not change in the soils during the experiment. In the agricultural soil total mercury was 2.7 20.2 mg Hg gÿ1 dry soil on d 0 and 2.6 2 0.1 mg Hg gÿ1 at the end of the experiment on d 15. During the same period, beech forest soil had 3.220.2 and 3.320.5 mg Hg gÿ1 dry soil, respectively. The background concentration of total mercury was similar in both soils (0.19 and 0.27 mg gÿ1 in agricultural and beech forest soil, respectively), the cause of the dierent concentrations of mercury after addition is not known. 3.2.2. Frequency of mercury resistant bacteria In the agricultural soil the amount of mercury resistance began to increase after the ®rst day of exposure. The frequency of mercury-resistant isolates increased rapidly reaching its peak on d 4 at 7.122% (Fig. 3B). A gradual decline followed until a plateau was reached between d 7 and d 11, where the % Hg resistance frequency remained, slightly above pre exposure quantities, for the duration of the experiment. No eect of the added mercury was observed on the frequency of mercury resistant bacteria in the beech forest soil with a constant fraction of approx. 0.004% of the total CFU growing on the mercury supplemented medium (Fig. 3B). 3.2.3. Bacterial diversity Diversity was evaluated by the number of bands visualised in DGGE gels after electrophoresis of PCR ampli®cation product obtained using primers speci®c to all eubacterial 16S rRNA genes. This analysis provides information on the diversity of the eubacterial community at large (i.e. without the need to culture soil bacteria) (Muyzer et al., 1993; Heuer and Smalla, 1997). Diversity of the indigenous bacteria in the two soils responded dierently to mercury exposure (Fig. 3C). In the agricultural soil diversity decreased rapidly in the ®rst 4 d of the experiment with the average number of 16S rDNA bands declining from 26 to 20. After this minimum the number of bands gradually increased for the rest of the experiment reaching the initial number on d 15. In the beech forest soil no eect of mercury exposure was observed on the number of 16S rDNA bands, i.e. diversity, which were about 27 for the duration of the experiment.
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4. Discussion This investigation showed that the mer-lux bioassay is a useful tool for quanti®cation of bioavailable mercury in soil leachate obtained from three dierent soils. In this study the water-leachable fraction of mercury is referred to as being potentially bioavailable. This is in agreement with Sterckeman et al. (1996) who found a strong correlation between water-extractable mercury and accumulation of mercury in plants. Bacterial activity in the soil is located in niches characterized by the availability of water, and the fact that Hg(II) needs to be in aqueous solution in order to be transported to the cytoplasm is supporting the assumption that bioavailable mercury in soil is water leachable. Results of experiments designed to optimize the merlux bioassay in soil leachate showed the biosensor response might be biased by the presence of soil particles. Forced ®ltration to remove soil particles, resulted in an underestimation of bioavailable mercury (Fig. 1A). Assays performed in water and ®ltrate spiked with 0.75 nM Hg(II) showed that this was due to mercury-binding ligands that were released during ®ltration, probably by pressure-induced cell lysis (Fig. 1B). This corresponds well with earlier ®ndings that dissolved organic carbon (DOC) considerably decreases biosensor response (Barkay et al., 1997). That removal of soil particles from leachates was needed to avoid underestimation of the concentration of bioavailable mercury was shown by comparing biosensor responses in assays performed on gravity ®ltered and centrifuged leachates with assays in raw untreated leachate (Fig. 2). Assays performed with the constitutive mutant showed that this was most probably due to shading by soil particles (Fig. 2B). Little is known about the toxicity of mercury associated with colloidal and ®ne particles. Since these fractions of the total mercury will be eliminated from the soil leachate by centrifugation eventual bioavailability of these will not be detected by this assay. The microcosm experiment showed that the microbial response to mercury observed as development of mercury-resistant bacteria and lowering of diversity was correlated to changes in concentrations of bioavailable mercury (Fig. 3). Both these responses indicate toxicity and are well documented eects of heavy metal contaminations in soil (Roane and Kellogg, 1996; Ranjard et al., 1997; Smit et al., 1998). The heavy metal concentration at which a bacterial response is elicited may vary greatly depending on the metal and the soil type (BaÊaÊth, 1989). In our study, the dierences in response of the two bacterial communities in the two soil types to dosing with equal amounts of total mercury, con®rms that bioavailable rather than total mercury is the factor controlling microbial responses. This is in good agreement with a study by Ran-
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jard et al. (1997) who after spiking four dierent soil types with the same mercury concentration found that after incubation the proportion of resistant bacteria varied from 0.4 to 36% . Many abiotic factors may aect the bioavailability of heavy metals in soil e.g. clay content, pH, dissolved organic carbon, root exudates (BaÊaÊth, 1989; Barkay et al., 1997; Giller et al., 1998). The main dierences in the soil variables (Table 1) between the agricultural and the beech forest soil is pH and ammonium (NH4) content. Soil pH is often found to have a large in¯uence on metal availability due to its strong eect on metal solubility and speciation. Thus, a decrease in pH usually results in increased availability (Giller et al., 1998). This seems not to be the case in this study where the pH in the beech forest soil was almost 3 units lower than that in agricultural soil (Table 1). The high ammonium content in the beech forest soil (Table 1) may be an important factor decreasing bioavailability of mercury. Experiments with varying the concentration of the dierent bioassay medium constituents have shown that the sensitivity of the assay is decreased by increasing the ammonia concentration (unpublished data). The results from beech forest soil shows that although bioavailable mercury was present in detectable concentrations in the soil, there was no enrichment of resistant bacteria or decline in community diversity (Fig. 3). These results suggest that bioavailable mercury at the detected concentration may not have been toxic to indigenous bacteria in beech forest soil. The bioassay requires the presence of bioavailable Hg(II) in the biosensor cytoplasm. This species of mercury is the substrate for mercury methylation by soil bacteria (Beckert et al., 1974). The fact that detectable bioavailable mercury in beech forest soil was not toxic suggests that a potential for mercury methylation exists at subtoxic concentrations of mercury. Acknowledgements L.D.R. is supported by ``Centre for biological processes in contaminated soil and sediment'' (www.biopro.dk) under The Danish Environmental research programme. The authors wish to thank Pia Kringelum and Inge E. Larsen for excellent technical assistance. References Barkay, T., Turner, R., Saouter, E., Horn, J., 1992. Mercury biotransformation and their potential for bioremediation of mercury contamination. Biodegradation 3, 147±159. Barkay, T., Gillman, M., Turner, R.R., 1997. Eects of dissolved organic carbon and salinity on bioavailability of mercury. Applied and Environmental Microbiology 63, 4267±4271. Barkay, T., Turner, R.R., Rasmussen, L.D., Kelly, C.A., Rudd, J.W.M., 1998. Luminescence facilitated detection of bioavailable
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