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a Department of Civil, Geological and Mining Engineering, École Polytechnique de Montréal, P.O. Box 6079, Station Centre-ville, Montreal, QC, Canada H3C 3A7
b NSERC Industrial Chair in Site Remediation and Management, Chemical Engineering Department, École Polytechnique de Montréal, P.O. Box 6079, Station Centre-ville, Montreal, QC, Canada H3C 3A7
* Corresponding author (gerald.zagury{at}polymtl.ca)
Received for publication March 10, 2004.
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ABSTRACT |
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 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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Abbreviations: CAP, chlor-alkali plant • CRM, certified reference material • F1, water-soluble fraction • F2, exchangeable fraction • F3, organic fraction • F4, residual fraction • SEP, sequential extraction procedure
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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The contamination of soils in the vicinity of a CAP is due to emissions containing elemental Hg (Hg0), which is almost insoluble and can react with metallic oxides and organic matter (Schuster, 1991; Stein et al., 1996). In fact, mercury has an extremely high affinity for organic matter and sulfur ligands (Wallschläger et al., 1998). Therefore, its mobilization is generally of minor importance in organic soils, although it can be reemitted in the atmosphere especially during periods of high temperature (Rule and Iwashchenko, 1998) or can be oxidized under acidic conditions and transformed into its bivalent form, Hg2+ (Hempel et al., 1995). Divalent Hg is a necessary precursor for the formation of compounds with increased solubility and mobility like HgCl2 and Hg(OH)2 or increased bioavailability such as methyl Hg or ethyl Hg. Therefore, as for other metals, Hg speciation is essential because it controls its solubility, volatility, reactivity, bioavailability, and finally, its toxicity (Stein et al., 1996). The toxicity of Hg is well known for some species (e.g., HgCl2, CH3HgCl) but it is much more difficult to assess in real samples from contaminated sites. Furthermore, differences in the solubility and volatility of the various species of Hg affect the concentration of Hg available to the exposure pathway. For example, the three principle routes of Hg uptake by vascular plants are: (i) through the roots from the soil solution as ionic Hg, (ii) through the stomata from the atmosphere as volatile Hg, and/or (iii) through foliar adsorption of divalent, reactive gaseous Hg and particulate Hg (Ericksen and Gustin, 2004). The determination of fractionation is an important step in the evaluation of the toxicity potential of Hg-contaminated soils from CAPs, and a strong need exists for appropriate solid-phase speciation schemes.
Published groups of methods for the determination of Hg speciation include sequential extraction procedures (SEPs) (Di Giulio and Ryan, 1987; Revis et al., 1989; Lechler et al., 1997; Wallschläger et al., 1998; Bloom et al., 2003), short extraction techniques (Tomiyasu et al., 2000), pyrolitic extractions usually called mercury-thermo-desorption techniques (MTD) (Biester and Nehrke, 1997; Biester and Scholz, 1997), a combination of MTD with a partial sequential extraction (leaching test) (Biester et al., 2002) or X-ray absorption spectroscopy (XAS) (Kim et al., 2000).
Despite some criticism (technique validation, operational definition, limited selectivity, and possible redistribution of metals) dealing with the interpretation of SEPs (Nirel and Morel, 1990), their use has continued to be recognized as a valuable tool, provided they are used with discrimination and caution (Tessier and Campbell, 1991; Bloom et al., 2003; Sladek and Gustin, 2003). Sequential extraction procedures yield pragmatic information on the possible behavior of Hg in the soil environment by using extractions with chemical reagents that destroy the binding agent between the metal and the soil solids.
Along with its strengths, each speciation technique and fractionation scheme also has its limitations; available SEPs for Hg are sometimes incomplete because they do not assess the volatile Hg fraction, nor were they tested on highly contaminated ambient soil samples (Di Giulio and Ryan, 1987; Wallschläger et al., 1998). The shorter published SEPs (Tomiyasu et al., 2000) were used to investigate Hg fractionation in contaminated sediments with low concentrations of Hg (0.086–3.46 mg kg–1). As for pyrolytic extractions, they have limitations because they do not give information about Hg forms in soil leachates (Biester and Scholz, 1997; Biester and Nehrke, 1997).
Although the inherent SEP limitations remain, here we present a novel Hg-specific solid-phase speciation scheme, developed to be easily implemented by most trace metal laboratories and tested on highly contaminated field-collected soil samples obtained from CAPs. The extraction scheme was also applied on a certified reference material (BCR CRM 580) and on three pure Hg compounds dispersed in kaolin for technique validation. The developed SEP consists of four steps for estimating the distinct fractions of Hg removed in specific environments (neutral, slightly acidic with exchangeable cations, alkaline with complexing extractants, strongly alkaline, and strongly acidic). In addition to the sequential extractions, total, volatile, and methyl Hg concentrations were determined on separate soil samples and on the CRM using single extractions. The proposed SEP was very precise and yielded satisfactory recovery percentages.
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MATERIALS AND METHODS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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The SEP and the single extractions were also applied on the Community Bureau of Reference BCR 580 CRM (estuarine sediment). This CRM was selected because it was subjected to the five-step fractionation scheme of Bloom et al. (2003) and to the extraction scheme (ethanol extraction and acid extraction) published by Han et al. (2003) allowing intercomparison. Furthermore, three pure standard Hg compounds (HgCl2 [Anachemia, Montreal, QC, Canada], Hg0 [Aldrich, St. Louis, MO], and HgS red [Aldrich]) separately dispersed into kaolin powder (Sigma) at concentrations ranging from 35 to 740 mg Hg kg–1, were extracted using the proposed SEP. Batches of 100 g of each pure Hg compound–kaolin mixture were prepared in a 500-mL polypropylene copolymerized (PPCO) screw cap bottle and agitated on a customized rotary agitator.
Soil Characterization
Soil samples were characterized for pH, water content, cation exchange capacity, chlorides, total sulfur, total carbon, total inorganic carbon, total volatile solids, and particle-size distribution. The pH was measured in deionized water using a solid to liquid ratio of 1:1 according to Method D 4972-95a (American Society for Testing and Materials, 1995) using an Accumet Model AR 25 pH meter and an Accumet 13-620-285 combination Ag/AgCl electrode (Fisher Scientific, Hampton, NH). Moisture content was determined at 45°C to minimize Hg volatilization during drying (Maserti and Ferrara, 1991; Sakamoto et al., 1995). Cation exchange capacity was determined using the sodium acetate method (pH 8.2) according to Chapman (1965), chlorides determination was performed on an aqueous soil extract by ion chromatography separation (Dionex [Sunnyvale, CA] Model DX-100), and electrochemical (conductivity) detection was performed according to Standard Method 4110 (Clesceri et al., 1998). Volatile solids were determined at 550°C according to Karam (1993). Total carbon and total sulfur were measured by combustion with an induction furnace (LECO, St. Joseph, MI). A phosphoric acid treatment followed by an infrared determination of CO2 evolved was performed to determine total inorganic carbon (Ministère de l'Environnement et de la Faune du Québec, 1996). Organic carbon was calculated by the difference between total carbon and total inorganic carbon. The samples were classified as gravel (>2 mm), sand (2 mm–75 µm), and silt and clay (<75 µm) according to Method D 2487-83 (American Society for Testing and Materials, 1985).
Total Mercury, Volatile Mercury, and Methyl Mercury
The total concentration of Hg in contaminated soils (n = 5) was determined following an acid digestion scheme adapted from Akagi and Nishimura (1991). One gram of soil was accurately weighed and placed in a 100-mL volumetric flask and 14 mL of 1 HNO3 to 5 H2SO4 to 1 HClO4 were added. The digestion was performed at 250°C on a hot plate (Corning [Corning, NY] PC 320 model) for 1 h, and allowed to cool to room temperature. The digest was filtered (0.45 µm) and diluted with deionized water to 100 mL. Total Hg was measured by cold vapor atomic absorption spectrometry (CVAAS) (CETAC Technologies [Omaha, NE] Model M-6000A). The detection limit for Hg in the liquid phase was 0.1 µg L–1. To investigate the assumption that most of the mercury is associated with the <2-mm fraction, total Hg was also measured in the gravel fraction (>2 mm) using the same procedure.
Accuracy and precision of the digestion procedure for total Hg analysis were also verified using the certified reference material BCR CRM 580 (n = 3). The reference value is 132 mg kg–1 with an uncertainty of 3 mg kg–1 (half-width of the 95% confidence interval of the mean). Our mean concentration (n = 3) was 129 mg kg–1 with a standard deviation of 2 mg kg–1, which is consistent with the certified value (Table 2). Mercury concentrations in procedure blanks and in all reagents were always below the detection limit.
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Methyl Hg was measured in contaminated soils (n = 4) and in the BCR CRM 580 (n = 4) by gas chromatography separation and cold vapor atomic fluorescence spectroscopy detection (GC–CVAFS) after aqueous-phase ethylation, according to Liang et al. (1994). The absolute detection limit was 0.6 pg as Hg. The certified value of the CRM is 75.5 ± 3.7 µg kg–1. Our average value was 73.1 µg kg–1 with a standard deviation of 1.5 µg kg–1, which is in agreement with the certified value.
Sequential Extraction Procedure
Following an extensive literature review covering the determination of Hg speciation in solid samples, a novel SEP was developed, based partially on the work of Di Giulio and Ryan (1987), Lechler et al. (1997), and Wallschläger et al. (1998). The proposed SEP provides differentiation of Hg compounds into the following four fractions: F1, water-soluble; F2, exchangeable (two alternative procedures were tested); F3, bound to organic matter; and F4, residual Hg.
In practical terms, the fractions can correspond to the following behaviors in natural conditions: F1, fraction of Hg that can be easily released, such as Hg immediately available following a simple leaching by rainwater; F2, potentially available Hg fraction under alkaline conditions (pH = 8.4) in the presence of a complexing agent (Alternative 1) or under acidic conditions (pH = 5) with exchangeable cations (Alternative 2); F3, Hg fraction that is available under acidic (pH = 3) or alkaline (pH = 13) conditions; and F4, Hg fraction considered to be weakly soluble and that can only be released by a strong attack of the matrix.
The sequential extractions were performed using 2 g of accurately weighed soil sample mixed with 20 mL of solvent in a 50-mL polypropylene centrifuge tube. The tubes were thoroughly shaken for 2 h at 20 ± 2°C using a mechanical wrist-action shaker (Burell Scientific [Pittsburgh, PA] Model 75). Between each of the successive extractions and rinses, the supernatant was obtained by centrifuging (Beckman [Fullerton, CA] Model J2-21) at 12000 x g for 15 to 25 min at 10°C. Rinsing steps consisted of washing the leached residues twice with deionized water (20 and 10 mL) for 15 min. The rinses were then added to the solvent extract from the same sample. The combined supernatant was analyzed for Hg by CVAAS. The solid residue was used in the next extraction step.
F1: Water-Soluble Compounds
The water-soluble fraction extractions were performed in at least triplicate samples, for each of the two alternative procedures. The extraction procedure was conducted as described above using deionized water as solvent. After the rinse steps, the residue was saved for the next extraction.
F2: Exchangeable Compounds
Panda et al. (1990) investigated three chemical extractants to assess the bioavailability of Hg (10% HNO3, 0.025 M NH4COOCH3–0.02 M EDTA and 0.1 or 0.05 M CaCl2) and they found that CaCl2 was the best predictor of bioavailable Hg. McLaughlin et al. (2000) also reported that CaCl2 and 0.5 M NH4Ac–EDTA were good predictors of phytoavailable metals in soils but they warned that rigorous comparisons with other procedures over a wide range of soil types were lacking. Therefore, in the present work, two chemical extractants were tested. The exchangeable fraction was extracted under alkaline conditions with a complexing agent (0.5 M NH4Ac–EDTA (pH = 8.4), (Alternative 1) and under slightly acidic conditions with 1 M CaCl2 (pH = 5) (Alternative 2). Each method was performed on three or four replicates while respecting the same working conditions as described above. To prepare the 0.5 M NH4Ac–EDTA, CH3COONH4 (38.54 g) and EDTA (186.12 g) were accurately weighted and separately dissolved in two 500-mL Pyrex beakers using a minimal amount of deionized water to dissolve the crystals. For the EDTA dissolution, the pH was adjusted to 8.0 with 28% (w/v) ammonium hydroxide. The two solutions were mixed into a standard volumetric flask (CH3COONH4 solution was added in first) and the volume made up to 1000 mL with deionized water.
F3: Organic Compounds
This fraction was separated by successive extractions using 0.2 M NaOH and CH3COOH 4% (v/v). After recuperation of the first extract (0.2 M NaOH), CH3COOH was added. Finally, two rinses with deionized water were performed (20 and 10 mL) and then added to the CH3COOH extract. The 0.2 M NaOH supernatant and the combined CH3COOH supernatant were analyzed for Hg by CVAAS. The solid residue was used in the next extraction step.
F4: Residual Compounds
Residual Hg was extracted by adding the same reagents as for total Hg to the soil pellet directly in the original 50-mL centrifuge tube. The sample was then transferred into a 100-mL standard volumetric flask. The digestion was performed using the same procedure as previously described for total Hg.
All laboratory ware used during the analytical procedures was cleaned sequentially with a phosphate-free detergent, soaked in 10% (v/v) nitric acid for 24 h, then in distilled water, and finally rinsed three times with deionized water (18.2 Mohms). Unless otherwise stated, all reagents were of analytical grade (ACS) or better. Statistical treatment of data was performed using STATISTICA software (5.1 version; Statsoft, 1997).
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RESULTS AND DISCUSSION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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Volatile Mercury
The highest concentrations of volatile Hg were found in Soils 1 and 3, representing 88 and 98% of total Hg, respectively (Table 2). The undisturbed surface Soil 2 had a lower concentration of volatile Hg (42 ± 19 mg kg–1). The results suggest the presence of very high concentrations of Hg0 or potentially volatile HgCl2 in Soils 1 and 3. Soil 3, in particular, seems to be the most hazardous from a storage or treatment perspective, with 11300 mg Hg kg–1 in the volatile form. Sladek and Gustin (2003) reported that pyrolytic extraction of Hg at a temperature of 80°C significantly extracted Hg from HgCl2–amended ground glass but even at 180°C, Hg0 was not completely extracted when small amounts of organic matter (0.4–3%) were present in the samples. Unfortunately, thermal extraction applied at 180°C for 2 d also leads to volatilization of HgCl2 and Hg bound to organic matter (Biester and Scholz, 1997; Bloom et al., 2003). In our study, thermal extraction of pure HgCl2 and pure Hg0 amended to kaolin powder showed that 81% of total Hg and 84% of total Hg, respectively, were extracted after 48 h at 180°C. Also, 20% of total Hg was found to be volatile in the CRM. Such levels of volatile Hg in soils are unusual, but not odd, because Hg is used in the elemental form as a cathode in caustic soda productions. Biester and Nehrke (1997) reported that on the average, 34% of all Hg occurred as Hg0 in a highly contaminated soil (1717 mg Hg kg–1) by emissions from a CAP. In a recent study (Biester et al., 2002), however, the same author reported undetectable Hg0 in somewhat contaminated soils from atmospheric deposition of Hg (<4.2 mg Hg kg–1) collected in the vicinity of CAPs. In summary, the results of both pyrolytic extraction of Hg at a temperature of 80°C and of thermal extraction at 180°C should definitely be taken as an overestimation of elemental Hg. This volatile Hg extraction step is, however, interesting for the evaluation of the potential environmental hazard caused by Hg-contaminated soils (for example, atmospheric Hg uptake by plants through their stomata).
Methyl Mercury
Methyl Hg concentrations (Table 2) were low (from 0.2 to 19.3 µg kg–1) compared with published soil values (from 670 to 10000 µg kg–1) used in screening, cleanup, or monitoring of contaminated sites (American Petroleum Institute, 2000). No significant correlation was found between total Hg and the methyl Hg fraction. The higher organic carbon content of Soil 2 (18.2 g kg–1) did not entail a high proportion of methyl Hg (0.00008% of total Hg). Hempel et al. (1995) and Bloom et al. (2003) reported methyl Hg concentrations of 400 µg kg–1 (<0.04% of total Hg) and 10 µg kg–1 (<0.0001% of total Hg), respectively, in polluted soils near chlor-alkali plants, stressing the site-dependent aspect of the methyl Hg fraction.
Fractionation of Pure Mercury Compounds and Certified Reference Material
To validate our solid-phase fractionation scheme, three pure Hg compounds dispersed in kaolin powder were extracted using the developed technique (Table 3). As would be expected, compounds with very low solubility such as HgS and with low solubility such as Hg0 were not significantly extracted by any of the solutions (F1 to F3) regardless of the alternative used for F2. However, between 99.0 and 99.9% of HgS and Hg0 were recovered during the extraction with the mixture of strong acids (1 HNO3 to 5 H2SO4 to 1 HClO4) corresponding with residual Hg (F4). These results are consistent with the observations reported by Bloom et al. (2003), who used a five-step fractionation scheme as follows: F1 (deionized water), F2 (0.01 M HCl + 0.1 M CH3COOH), F3 (1 M KOH), F4 (12 M HNO3), F5 (aqua regia). In our study, however, water-soluble HgCl2 (F1) ranged between 61.9 and 84.8% of total Hg. The remainder of this species was leached during the extraction with NH4Ac–EDTA (8.1%) and CaCl2 (19.1%) and was completely recovered following extraction with strong acids. This quite surprising result can be attributed to homogeneity problems encountered during the preparation of the pure HgCl2–amended kaolin powder.
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Sequential extraction procedure results indicate that the sum of F1 (water-soluble), F2 (exchangeable), and F3 (organic) represents low percentages of total Hg (below 42, 12.5, and 5% [Alternative 1] and below 14, 4, and 3% [Alternative 2] for Soils 1, 2, and 3, respectively). For both alternatives, the most important contribution to the sum of these three fractions was the second extraction (F2) with results varying between 11.1 and 39.6% (Soil 1), 3.1 and 12.0% (Soil 2), and 2.1 and 4.4% (Soil 3) of total Hg. In contrast, F3 (Hg associated to organic matter) was very low, with percentages of total Hg ranging from 0.1 to 2.0%, regardless of the extraction scheme. These results are consistent with the low organic carbon content of the three study soils. The Hg speciation pattern obtained in this study is, however, very different from the one found in much less contaminated organic soils polluted by atmospheric deposition of Hg around CAPs (<4.2 mg Hg kg–1), where Hg0 was undetectable and Hg was predominately bound to organic matter (Biester et al., 2002).
Although the water-soluble fraction (F1) only accounts for at most 1.1% of total Hg (Soil 1), this fraction is recognized as very important from an environmental risk point of view due to its easy availability in environmental weathering conditions (Wallschläger et al., 1998; Bloom et al., 2003). Moreover, even these small percentages should be treated with caution because they represent in fact very significant Hg concentrations. In the case of Soil 3, for example, although the water-soluble fraction is lower than 0.5% of total Hg, it still represents more than 50 mg Hg kg–1. Such a concentration exceeds Quebec's background concentration in soils (Level A) and the ecological soil criteria of Finland and Sweden (De Vries and Bakker, 1998) by a factor of 250. One explanation for the varying values obtained for the mobile fractions (F1 and F2) in the different soils may be in the ratio of chlorides to total Hg, which was positively correlated with exchangeable Hg (F2) (r = 0.99, P = 0.02 for the SEP Alternative 1, and r = 0.99, P = 0.09 for the SEP Alternative 2) and with the sum (F1 + F2) (r = 0.99, P = 0.03 for the SEP Alternative 1, and r = 0.98, P = 0.11 for the SEP Alternative 2). This suggests that Hg mobility increases with the ratio of chlorides to total Hg rather than with the chloride level in soils, as previously reported (Schuster, 1991; Yin et al., 1996). The number of samples is, however, too limited to provide broad geochemical interpretation, but the observed correlations reflect the needs of future research in this area.
In all soils, the largest Hg proportion was found within the residual fraction, which ranged from 65.6 to 70.8% (Soil 3) to 73.5 to 116% (Soil 1) of total Hg. In naturally occurring conditions, the residual fraction (F4) represents the least available form of Hg, depending on the matrix under study and the source of contamination (Wallschläger et al., 1998; Bloom et al., 2003). In the case of Soils 1 and 3, however, the residual fraction should be interpreted in a different way due to the presence of very high concentrations of volatile Hg. Therefore, in these two soils, a high proportion of volatile Hg and hence, Hg0, was present in the residual fraction. Consequently, a distinction should be made between availability in terms of solubility and availability in terms of volatility. In other words, the various routes of Hg uptake (from the soil solution or atmospheric Hg) and the various receptors should be kept in mind when interpreting fractionation results. There is also reason to believe that for all soil samples, trace amounts of elemental Hg are present in every fraction extracted (Biester and Scholz, 1997) because it has a water solubility of approximately 50 µg L–1 (Bloom et al., 2003). In fact, Hg0–amended kaolin powder fractionation (Table 3) showed that no more than 0.1% of total elemental Hg was extracted in F1, F2, and F3 with SEP Alternative 1 (0.5 M NH4Ac–EDTA [pH = 8.4] for F2), whereas 0.1, 0.8, and 0.0% of total elemental Hg was extracted in F1, F2, and F3, respectively, with SEP Alternative 2 (1 M CaCl2 [pH = 5] for F2). The different solubility of Hg0 in F2 is possibly attributed to the presence of chlorides and to the different pH of the extract. Furthermore, the relatively lower recovery percentages for the SEPs conducted on Soil 3 (70 ± 5% and 73 ± 4%) suggest the loss of volatile Hg species during the manipulations. Volatilization of Hg0 was given as a possible explanation by Wallschläger et al. (1998) when they reported low average recoveries (66 ± 26% for floodplain soils and 44 ± 26% for sediments) following their five-step extraction procedure.
Table 4 reveals that the exchangeable fraction obtained with 0.5 M NH4Ac–EDTA was significantly higher than that obtained with 1 M CaCl2. In the presence of alkaline conditions and EDTA, the proportion of Hg extracted was 3.5, 4, and 2 times higher for Soils 1, 2, and 3, respectively, than the proportion extracted in the presence of slightly acidic conditions and calcium chloride. The lower difference in the proportions of exchangeable Hg extracted in Soil 3 might be explained by its higher carbonate content. Apparently, the only difference between the two alternative SEPs lies in the quantity of Hg extracted in F2. That being said, after F2 extraction, F3 underwent a slight change whereas F4 underwent a significant increase especially in Soils 1 and 2. This change affected the sum of fractions and consequently the recovery percentages.
Globally, Soil 1 showed the highest percentages of mobile (F1 + F2) Hg (12 or 42% of total Hg depending on the alternative). Once again, even though F1 appears to be relatively low in terms of percentages (about 1%), it nevertheless represents a significant portion in terms of concentration (about 5 mg kg–1). This concentration crosses the limit that separates a safe soil for agricultural use from an unsafe one according to Panda et al. (1992). Moreover, its grain-size distribution (sandy soil) entails a high hydraulic conductivity. Hence, mobile Hg in this soil could easily migrate to deeper soil layers and disperse via pore water.
Soil 2 had the lowest concentration of total Hg (295 ± 18 mg kg–1) and appeared to be much less dangerous in terms of volatility. This may be due to possible Hg0 re-emission. The Hg from chlor-alkali plant-contaminated soils volatilizes easily, especially during seasons with elevated temperatures (Rule and Iwashchenko, 1998). The amount of Hg that may undergo solubilization in the short and medium term (F1 + F2) was relatively lower in this soil (9.5 or 35.5 mg kg–1 at the most) and most Hg extracted was residual (extracted with a mixture of HNO3, H2SO4, and HClO4) under the form of nonvolatile weakly available Hg (most likely HgS). A long-term contamination and a lower chloride concentration (HgCl2 has approximately a water solubility of 69000 mg L–1) could explain the extremely small fractions of water-soluble Hg in this soil (0.1% compared with 0.7–1.1% and 0.4–0.5% for Soils 1 and 3, respectively). Nevertheless, a portion of the Hg found in this soil remains potentially phytoavailable through root uptake from the soil solution.
Soil 3 was extremely contaminated (11500 ± 500 mg kg–1), with 98% of total Hg in the volatile form. According to the SEP results, the proportion of easily leachable Hg (F1 + F2) was the lowest for this soil (2.6–4.8% of total Hg), while the majority of Hg was in the residual fraction, suggesting relatively lower potential for Hg mobility via pore water. However, as previously mentioned, such percentages should be treated carefully because total Hg concentration is very high in Soil 3. As a consequence, the water-soluble and exchangeable fractions actually represent important Hg concentrations (300–557 mg kg–1). Furthermore, even though the concentration of methyl Hg is relatively low (19.3 ± 1.5 µg kg–1), compared with values ranging from 10 to 9850 µg kg–1 in various contaminated soils, sediments, and tailings (Bloom et al., 2003), this species is still potentially toxic at such a concentration primarily due to its bioconcentration potential.
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CONCLUSIONS |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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The developed extraction procedure distinguished water-soluble (F1), exchangeable (F2) (two solutions were tested), organic (F3), and residual (F4) Hg. Both extraction alternatives for F2 gave the same Hg fractionation pattern, but higher percentages in the exchangeable fraction were obtained using 0.5 M NH4Ac–EDTA. Sequential extractions indicated that most of the Hg was associated with F4; F4 is the least available form of Hg based on its potential solubilization. In Soils 1 and 3, however, a high proportion of volatile Hg was present in the residual fraction.
Soil 1 showed the highest percentages of mobile Hg whereas mobile Hg percentages were much lower in the two other soils. The mobile Hg (F1 + F2) and the exchangeable Hg (F2) fractions were strongly and positively correlated with the chlorides to total Hg ratio. In all soils, Hg associated with organic matter (F3) was very low. These results are different from previously published observations but were in good agreement with the low organic carbon content of the study soils. Nevertheless, it should be noted that for Soil 3, the low percentages of water-soluble, exchangeable, and organic-bound Hg represent in fact very significant concentrations due to its extremely high total Hg content.
The developed SEP was very precise for all Hg-contaminated soils tested with satisfactory recovery percentages ranging from 70 to 130% and from 93 to 98% for the CRM. Knowledge of Hg fractionation in soils impacted by direct emissions of Hg from CAPs yields information about the potential Hg mobility and environmental impact but is also important in defining the decontamination strategy if one is needed. The very high total Hg concentrations in all study soils, as well as the results of Hg fractionation and speciation, suggest a potential ecotoxicity via the soil solution and the atmosphere. To confirm the likely uptake of the various Hg species by organisms and plants, bioassays were also performed on the Hg-contaminated soils using earthworm (Eisenia andrei) and barley (Hordeum vulgare L.). These results are expected to be published in the near future.
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ACKNOWLEDGMENTS |
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REFERENCES |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONCONCLUSIONSREFERENCES |
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ki, and H. Akagi. 2000. Mercury contamination in the Yatsushiro Sea, south-western Japan: Spatial variations of mercury in sediment. Sci. Total Environ. 257:121–132.[Medline]Related articles in JEQ:
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