Environ. Sci. Technol. 2008, 42, 4378–4383
Solubility and Toxicity of Antimony Trioxide (Sb2O3) in Soil K O E N O O R T S , ‡ E R I K S M O L D E R S , * ,† FIEN DEGRYSE,† JURGEN BUEKERS,† ´ ,§ GEERT CORNELIS,| GABRIEL GASCO AND JELLE MERTENS† Division Soil and Water Management, Katholieke Universiteit Leuven, Kasteelpark Arenberg 20, 3001 Heverlee, Belgium, EURAS - ARCADIS Belgium, Kortrijksesteenweg 302, 9000 Gent, Belgium, Departamento de Edafolog´ia, ´ ´ ETSI Agronomos, Universidad Politecnica de Madrid, Avda Complutense s/n, 28040 Madrid, Spain, and Laboratory of Applied Physical Chemistry and Environmental Technology, Katholieke Universiteit Leuven, De Croylaan 46, 3001 Heverlee, Belgium
Received December 7, 2007. Revised manuscript received March 12, 2008. Accepted March 19, 2008.
Antimony trioxide (Sb2O3) is a widely used chemical that can be emitted to soil. The fate and toxicity of this poorly soluble compound in soil is insufficiently known. A silt-loam soil (pH 7.0, background 0.005 mmol Sb kg-1) was amended with Sb2O3 at various concentrations. More than 70% of Sb in soil solution was present as Sb(V) (antimonate) within 2 days. The soil solution Sb concentrations gradually increased between 2 and 35 days after Sb2O3 amendment but were always below that of soils amended with the more soluble SbCl3 at the lower Sb concentrations. The soil solution Sb concentrations infreshlyamendedSbCl3 soils(7daysequilibration)wereequivalent to those in Sb2O3-amended soils equilibrated for 5 years at equivalent total soil Sb. Our data indicate that the Sb solubility in this soil was controlled by a combination of sorption on the soil surface, Sb precipitation at the higher doses, and slow dissolution of Sb2O3, the latter being modeled with a half-life ranging between 50 and 250 days. Toxicity of Sb to plant growth (root elongation of barley, shoot biomass of lettuce) or to nitrification was found in soil equilibrated with Sb2O3 (up to 82 mmol Sb kg-1) for 31 weeks with 10% inhibition values at soil solution Sb concentrations of 110 µM Sb or above. These concentrations are equivalent to 4.2 mmol Sb per kg soil (510 mg Sb kg-1) at complete dissolution of Sb2O3 in this soil. No toxicity to plant growth or nitrification was evident in toxicity tests starting one week after soil amendment with Sb2O3, whereas clear toxicity was found in a similar test using SbCl3. However, these effects were confounded by a decrease in pH and an increase in salinity. It is concluded that the Sb(V) toxicity thresholds are over 100-fold larger than background concentrations in soil and that care must be taken to interpret toxicity data of soluble Sb(III) forms due to confounding factors. * Corresponding author e-mail:
[email protected]; tel: +32 16329677; fax: +32 16321997. † Division Soil and Water Management, Katholieke Universiteit Leuven. ‡ EURAS - ARCADIS Belgium. § Universidad Polite´cnica de Madrid. | Laboratory of Applied Physical Chemistry and Environmental Technology, Katholieke Universiteit Leuven. 4378
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Introduction Antimony (Sb) is a naturally occurring trace element with increasing industrial use. Ambient Sb concentrations in European soils range from 0.1 to 1.9 mg kg-1 (or 0.001-0.016 mmol Sb kg-1, 10-90th percentiles; (1)). Antimony is mainly used as antimony trioxide (Sb2O3), especially in flame retardants, but also as a catalyst in plastics, as pigment in paints, or in production of glassware. Metallic Sb is used to harden lead alloys for the production of batteries and in ammunition. The annual world production of Sb was about 1.4 × 106 tons in 1999 (2). Antimony compounds are released into the atmosphere in the form of oxides during the incineration of waste, fossil fuel combustion, and during the smelting of metals. Other sources of contamination include road traffic (dust from brake linings and tires), shooting ranges (Sb in ammunition), and older battery producing plants. Soil Sb concentrations up to 130 mmol Sb kg-1 have been found in mining areas (2–5). Water-extractable Sb in mining soils is typically below 2% of total Sb in soil, corresponding to a concentration range of >0.001 to 0.12 mM Sb in water extracts (5–8). In aerobic soils, Sb(V) is the dominant redox species. Both metallic and trivalent Sb emissions in soil are oxidized to pentavalent Sb(OH)6- which is the dominant form observed with EXAFS and XANES spectroscopy in soils contaminated by smelters and shooting ranges (9, 10). Pentavalent Sb is also the dominant Sb species in surface water (8, 11). Trivalent Sb occurs as the neutral Sb(OH)3 at the environmentally relevant pH range (11) but readily oxidizes to Sb(V) in the presence of amorphous Fe and Mn oxides (12, 13). The Sb(V) in soil is mainly associated with Fe (hydr)oxides (8, 10, 14). Nakamaru et al. (15) reported solid-liquid concentration ratios (Kd) for Sb (added as SbCl3) in 110 Japanese agricultural soils after 7 days of equilibration in a 1:10 water extract. For 12 soils, the equilibrium solution was passed over an anion exchanger, and almost all Sb was retained, indicating that Sb in solution was present as the pentavalent Sb(OH)6- anion. The Kd values ranged from 1 to 2065 L kg-1 with a mean of 62 L kg-1. The Kd decreased with increasing pH and increasing phosphate concentration, suggesting sorption on oxyhydroxides. In addition to sorption on the solid phase, precipitation of Ca-antimonate at elevated total Sb may control solution concentrations of Sb in soil (8) or in waste material (16). Despite its wide use, there is still a general lack of data on Sb toxicity in the terrestrial environment. Antimony is a nonessential element and Sb(III) is reported to be about ten times more toxic in solution than Sb(V) (2). Flynn et al. (7) exposed two metabolic lux-marked biosensors to a range of Sb(III) (added as KSbO-tartrate) or Sb(V) (added as KSb(OH)6) concentrations in solution. Pentavalent Sb did not give bioluminescence inhibition up to a concentration of 0.8 mM Sb, while the Sb(III) concentrations yielding 50% inhibition (EC50) ranged between 0.006 and 0.05 mM Sb. Adverse effects attributed to Sb have not yet been identified in field studies where Sb concentrations ranged up to 130 mmol Sb kg-1 (3–7). Soil toxicity thresholds (EC50 values) identified in laboratory toxicity tests range from 0.5 to >8 mmol Sb kg-1 depending on the endpoint, soil, and Sb form tested (17–20). Different forms of Sb (Sb2(SO4)3, KSbO-tartrate, Sb2S3, Sb2S5, SbCl3, KSb(OH)6) were tested in these studies. None of them related toxicity of Sb with its (redox) speciation, and therefore comparison among these studies is difficult. In addition, none of these studies checked for confounding factors associated with the Sb-application such as pH changes and effects of 10.1021/es703061t CCC: $40.75
2008 American Chemical Society
Published on Web 05/09/2008
the anion, a recurrent issue when using metal salts to predict toxicity of metals in the environments (21, 22). Moreover, no information on toxicity of Sb2O3 in soils was found, despite its important industrial applications. The objective of this work was to quantify the toxicity of Sb2O3 in soil in relation to Sb solubility (soil solution concentration) and redox speciation. The solubility of Sb2O3 in soil was compared with that of a readily soluble compound, SbCl3, to obtain an estimate of the transformation rate of Sb2O3. Toxicity of Sb to plants and the nitrification process in soil was assessed at 6 months after Sb2O3 application with standard toxicity tests. A toxicity test with Sb2O3 after shortterm equilibration was referenced to SbCl3, which has been used in most laboratory toxicity tests with aquatic or sediment organisms. A salt control (CaCl2) test was included in that latter test to evaluate the effect of ionic strength, which is a confounding factor when deriving metal toxicity thresholds using metal salts (23).
Materials and Methods Four experiments were performed. Experiments 1 and 2 were designed to estimate the Sb2O3 transformation kinetics, and experiments 3 and 4 assessed Sb toxicity either after shortterm (experiment 3) or long-term (experiment 4) equilibration of Sb2O3 in soil. Soil and Soil Treatment. The soil used for all tests is an uncontaminated soil from an agricultural field (“Ter Munck”, Heverlee, Belgium). This soil is a Haplic Luvisol with a siltloam texture (14% clay, 76% silt), 0.93% organic carbon, a cation exchange capacity of 9.8 cmolc kg-1, and oxalateextractable Fe and Al concentrations of 2000 and 540 mg kg-1, respectively. The soil has a pH (0.01 M CaCl2) of 7.0 and contains 0.6 mg Sb kg-1 () 0.005 mmol Sb kg-1). After sampling, the soil was air-dried at 25 °C, sieved through 2 mm (unless stated otherwise), and mixed. In experiment 1, the dynamics of Sb2O3 reactions in soil was assessed based on a comparison of soil solution Sb concentrations between SbCl3 and Sb2O3 amended soils, as monitored between 2 and 35 days after amendment. Air-dry soil was amended with five concentrations of Sb2O3 or SbCl3 at 0-82 mmol Sb kg-1, added to the soil as a suspension. Additional deionized water was also added to adjust the soil moisture content to 29% (corresponding to a suction of 100 cm water, pF ) 2.0). All soils were mixed after amendments, and control samples (no Sb added) were treated in the same way as Sb-amended ones. Soils were incubated at 20 °C in darkness. Soil solution was extracted in triplicate regularly between 2 and 35 days after Sb amendment by centrifugation (3000 RCF, 15 min), filtered (0.45 µm), and analyzed. Soil pH (0.01 M CaCl2) of the incubated samples was determined in duplicate. In experiment 2, soil solution Sb concentration was measured in the soil amended with Sb2O3 (control, 0.08, 0.41, and 2.1 mmol Sb kg-1) 5 years before sampling. The Sb2O3 was mixed with white sand at various doses and 1 kg of Sb2O3 enriched sand was mixed with 40 kg of soil in a concrete mixer for 15 min. The amended soil was transferred to plastic containers 20 cm high with free drainage and aged outdoors. Soil samples were taken at 5 years after amendment and samples were air-dried and sieved through 2 mm. Soil solution was extracted after rewetting the soil to a moisture content at pF ) 2.0 and equilibrating at 20 °C for 35 days. The 5 years aging had not affected any of the soil properties measured, except for soil pH which was 7.5 in the aged samples, compared to 7.0 in the samples used for all other experiments. In experiment 3, the soil was amended with either Sb2O3 or SbCl3 at eight concentrations (0-16 mmol Sb kg-1) or with CaCl2 to identical total Cl concentrations as in the SbCl3
treated samples. Amending and mixing was done as in experiment 1. Soils were left to equilibrate for 1 week at 20 °C before starting the bioassays. At 4 weeks after amendment, the soil solution was extracted from samples that were not used for the bioassays and to which no NH4+-substrate (for nitrification test) or fertilizer solution (for plant growth test) was added. The electrical conductivity of all samples was measured in a saturation paste extract and soil pH was measured in a 0.01 M CaCl2 extract. For the nitrification test, triplicate 100 g subsamples were amended with 100 mg NH4-N per kg fresh soil using a stock solution containing 47 mg (NH4)2SO4 mL-1. The potential nitrification rate (PNR, mg NO3-N kg-1 day-1) was calculated from the net increase of soil nitrate in 7 days after the addition of NH4+ (23). Soil nitrate concentration was measured colorimetrically in a centrifuged soil extract (1 M KCl, soil/solution ratio ) 2.5, 2 h end-over-end shaking). The lettuce shoot yield test was based on the ISO 11269-2 protocol (24). Twenty lettuce seeds (Lactuca sativa cv. Pontiac) were sown in 3 replicate pots (600 mL, 10 cm diameter, 750 g soil per pot) for each soil treatment. Before sowing, 25 mg N and 6 mg P per kg of moist soil was added as fertilizer solution (18 mM NH4NO3, 18 mM KNO3, 6 mM KH2PO4). The soil surface was covered with a thin layer of polyethylene beads to reduce evaporation. The pots were placed in a growth cabinet (Weiss, 18′ SP/+5 Ju-Pa) with a 12 h/12 h day/night cycle. Air temperature in the cabinet was 20 °C (day) or 16 °C (night), and the relative humidity was 75% throughout. Light intensity at canopy height was 650 µmol photons m-2 s-1. Water loss from the pots was restored daily with deionized water. Following emergence, seedlings were thinned to 5 plants per pot. At 24 days after sowing, shoots were cut just above the soil surface and dried at 70 °C for at least 48 h, and the dry matter yield was recorded. In experiment 4, soil was mixed with Sb2O3 at six concentrations ranging from 0 to 82 mmol Sb kg-1. The soil was placed outside in plastic pots with free drainage for 31 weeks. Afterward, aged soils were air-dried and sieved (4 mm) and rewetted to 23% moisture content. Two plant growth assays and a nitrification test were performed. The lettuce shoot yield was determined as in experiment 3 except for the following details. The soils were fertilized with 50 mg P kg-1 as KH2PO4 and 100 mg N kg-1 as Ca(NO3)2 · 4H2O, and the plant assays started at 6 days after soil rewetting. Each soil sample was divided into 4 pots (9.5 cm diameter, 500 g moist soil per pot). Plants were harvested at 17 days after sowing and the fresh shoot biomass was recorded. Soil solutions were extracted in quadruplicate from the soil samples on which lettuce was grown. A barley root elongation assay was performed based on ISO 11269-1 (25). Barley (summer barley, Hordeum vulgare cv. Mauritia) was pregerminated (24 h, 25 °C, absence of light) and 4 seeds were planted in triplicate pots (ca. 9 cm depth, 100 g moist soil per pot) for each treatment. Pregerminated barley seeds with radicle smaller than 2 mm in length were planted. The plants were grown as in experiment 3 and water loss of the pots was restored daily with deionized water. After 5 days, intact roots were washed out of the soil and the length of the longest root on each plant was recorded. The average root length of each replicate pot was determined. The nitrification assay was performed as in experiment 3 except that the test started at 3 days after soil rewetting and that duplicate samples were used. Measurements. Solution compositions from experiments 1, 2, and 3 were analyzed by inductively coupled plasma optical emission spectroscopy (ICP-OES, PerkinElmer Optima 3300 DV). The Sb emission line at 217.6 nm was selected and the quantification limit was 10 µg L-1 (8 × 10-8 M). The Sb concentration in soil solutions from experiment 4 were measured using ICP-MS (Yokogawa HP 4500 series), with a VOL. 42, NO. 12, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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detection limit of 0.7 µg Sb/L (6 × 10-9M). Redox speciation of Sb in solutions from experiment 1 was measured in 2-25fold diluted soil solutions isolated from soil amended with 16 mmol Sb kg-1 at 2, 7, and 35 days (SbCl3) or 2 and 35 days (Sb2O3) after amendment. The solutions isolated after 35 days were stored at 4 °C in the presence of citrate pending the speciation analysis. The other samples were assayed directly after the soil solution isolation. The Sb speciation in solution was measured by coupling a Dionex ICS-2000 system to a Thermo X-series I ICP-MS. The IC system as well as the coupling to the ICP-MS was completely composed of PEEK tubing. The method of Krachler and Emons (26) was applied, originally intended for measuring low concentrations of Sb in urine samples. As Sb(III) is uncharged at ambient pH, EDTA is added to the mobile phase and Sb(III) elutes as a negatively charged Sb(III)-EDTA complex on a strong anion exchange column, a 4 mm guard, and analytical Hamilton PRP-X-100 column. A higher EDTA concentration (20 mM) was used relative to other methods described by Krachler and Emons (26) which ensured all possible Sb(III)-Cl complexes to be transformed into EDTA complexes and improved Sb(III) quantification. However, higher detection limits (three times the standard deviation of blanks), were obtained in this study (11 µg L-1 for Sb(III), 5 µg L-1 for Sb(V)) than by Krachler and Emons (26) (0.02 and 0.008 µg L-1, respectively) who applied HPLC coupling to hydride generation ICP-MS (HPLC-HG-ICP-MS). The high EDTA concentration in the mobile phase usually generates a high baseline value and consequently a relatively lower signal-to-noise ratio. Total metal concentrations in soils were determined by hot aqua regia digestion in a closed Teflon vessel (2 h, 140 °C for 13 min in microwave, Milestone Microwave Laboratory Systems), followed by determination with ICP-OES (experiments 1-3) or ICP-MS (Yokogawa HP 4500 series; experiment 4). Standard addition tests (Sb standard added to soil before digesting) yielded a recovery of 92 ( 3%. Lower and variable recoveries were found with open digestion vessels which are attributed to volatilization of Sb from HCl containing solutions (E. Steinnes, personal communication). Therefore, all digestions were made in closed systems. Soil total Sb concentrations were, on average, within 15% of the nominal values () background concentration + added dose) in soils that were not aged outdoors (no leaching). The soils from experiment 2 (outdoor aged soils) had lost about 50% of the added Sb during the 5-year aging period through leaching. Soils of experiment 4 (31 weeks outdoor aging) had lost less than 10% of the added Sb. Data Analysis. The ionic strength of each soil solution extract was calculated from the concentrations of cations, assuming that all anions were monovalent. Activity coefficients were calculated from the ionic strength based on the Debye-Hu ¨ ckel equation. The data from the bioassays were statistically examined through the application of a log-logistic dose-response model fitted with the Marquardt method (SAS v 8.02). The “dose” in this model is the added Sb (total Sb minus background or soil solution Sb minus background), with the dose in the control soil attributed a small value (0.001 mmol kg-1). These curves predict the concentrations at which 10 or 50% inhibition is found (EC10 and EC50). The highest no observed effect concentration (NOEC) was determined by ANOVA with a Duncan test.
TABLE 1. Concentrations of Sb(III) and Sb(V) in Soil Solution at Different Times after the Soils Were Amended with 16.4 mmol Sb kg-1 as SbCl3 or Sb2O3 (Standard Deviation in Brackets) soil solution concentration (µmol L-1) Sb form
incubation time (days)
Sb(III)
Sb(V)
SbCl3
2 7 35 2 35
107 (15) 82 62 (27->82)
80 150 150
110 (70-180) nd 240 (40->440)
320 (26->440) >440 340 (16->440)
a
nd ) not determinable (poor curve fitting).
solubility of Sb is controlled by sorption, presumably on Fe hydroxides (13). We estimated the solid-liquid distribution coefficient (Kd), i.e., the ratio of sorbed Sb concentration to solution concentration, from this linear part of the curve (Figure 3), which resulted in an estimated Kd of 38 L kg-1. This value is within the range of Kd values of Sb in soils reported in the literature (1-2065 L kg-1 with a mean of 62 L kg-1; (16)). The sorbed Sb in aerobic soils is most likely dominated by Sb(V) (see Introduction). No solid phase speciation was carried out in our study, so we cannot exclude that a small fraction of Sb on the solid phase was present as Sb(III), which sorbs stronger on the solid phase than Sb(V) (13). However, the sorption isotherm is linear (for Sb concentrations below 3 mmol kg-1), which would imply that the ratio Sb(III)/Sb(V) was the same at all concentrations and all sampling times. The plateau in soil solution Sb concentration suggests that Sb concentration was controlled by precipitation reactions. Johnson et al. (8) suggested that Sb concentration in leaching solutions from shooting range soils was controlled by Ca-antimonate precipitation, and reported a solubility product, Ksp, of 10-12.55 (Ksp ) (Ca2+)(Sb(OH)6-)2). This value predicts a maximal solubility of Sb(V) of about 11 µM at a Ca concentration of 5 mM, while we observed Sb concentrations of 330 µM or more. Our data suggest a solubility product (ion activity based; activities estimated by multiplying total solution concentrations by the activity coefficient) of Ksp ) 10-9.8. Our estimate for the solubility product is almost 3 log units larger than the value reported by Johnson et al. (8). Complexation of antimonate in soil solution could partly explain this discrepancy, as this would result in smaller antimonate activities than estimated. Complex formation through ion pairing of antimonate with cations such as Ca2+, Mg2+, and K+ (13) could thus have resulted in an overestimation of the solubility product, especially since Ca concentrations reached large values in the SbCl3-amended soils (Figure 2). Also complexation with dissolved organic matter might occur, but is likely of little importance (11). To obtain more evidence that the plateau in Sb solubility at the high SbCl3 doses was due to precipitation, we performed a small batch experiment. The soil solution from an Sb2O3-amended soil (pH 6.9) was diluted with distilled water or a Ca(NO3)2 solution to obtain Ca concentrations between 0.25 and 374 mM (40 µM Sb in all solutions). The solutions were left for one week at room temperature, and a visible precipitate formed in the solutions with high Ca concentration. The solutions of samples containing precipitates were filtered (5 mmol Sb kg-1), predicted equilibrium solution concentrations of Sb are larger in Sb2O3-amended soils than in SbCl3amended soils. This can be explained by the pronounced increase in soluble Ca in SbCl3-amended soils which promotes the precipitation of Ca[Sb(OH)6]2. Toxicity. Despite the wide use of Sb2O3 in industry, we did not find terrestrial toxicity data of this specific Sb compound. In our short-term equilibration experiment 3, contamination of the soil with Sb2O3 did not result in toxic effects on any of the endpoints. Plant shoot Sb concentrations increased in response to Sb2O3 addition, up to about 0.1 mmol Sb kg-1 dry weight, but no toxicity was detectable (Figure S2). In contrast, amending the soil with SbCl3 up to the same total Sb concentration (16 mmol Sb kg-1 soil) clearly resulted in toxic effects on all endpoints. However, these effects were partly confounded by salt and, potentially, pH effects. It is clear that toxicity tests with Sb(III) salts have little relevance for predicting the risks of Sb2O3 emissions to soil. In the long-term equilibration (31 weeks) experiments, toxicity was detected in the Sb2O3-amended soils, in which there were no confounding effects of pH or ionic strength. It might be argued that tests with soluble Sb(V) salts such as KSb(OH)6 could be used as the basis for assessing the long-term risks of Sb2O3 in aerobic soils where Sb(V) dominates. We found only one such study for terrestrial organisms but that study was performed with rice grown in submerged soils where reduction to Sb(III) might occur (18). Unfortunately, the Sb speciation was not measured. Our long-term equilibration test (31 weeks) was still not sufficiently long to yield equilibrium soil solution concentrations (i.e., complete transformation of Sb2O3) since these were still about a factor of 2 below the equilibrium line shown in Figure 3. This equilibrium line predicts that the lowest soil
solution EC10 (110 µM) corresponds to a total soil Sb EC10 value at equilibrium of 4.2 mmol Sb kg-1 (i.e., 510 mg Sb kg-1). This indicates that Sb toxicity thresholds in fully equilibrated aerobic soils (>1 year equilibration required) would be more than 2 orders of magnitude above the ambient Sb concentration range (0.1-1.9 mg Sb kg-1).
Acknowledgments We thank the International Antimony Oxide Industry Association (IAOIA) for funding this research. The study can not be freely used to comply with regulatory requirements like REACH without the formal agreement of IAOIA. A. Ruttens from LUC Diepenbeek is acknowledged for the setup and sampling of experiment 2. Author F.D. thanks the Fund for Scientific Research (FWO-Vlaanderen) for a postdoctoral fellowship.
Supporting Information Available Toxicity data and plant Sb concentrations from experiment 3. This information is available free of charge via the Internet at http://pubs.acs.org.
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(12)
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Literature Cited (1) Salminen, R. Ed. Geochemical Atlas of Europe. Part 1: Background Information, Methodology and Maps; Espoo: Geological Survey of Finland, 2005. (2) Filella, M.; Belzile, N.; Chen, Y. W. Antimony in the environment: a review focused on natural waters. I. Occurrence. Earth Sci. Rev. 2002, 57, 125–176. (3) Hammel, W.; Debus, R.; Steubing, L. Mobility of antimony in soil and its availability to plants. Chemosphere. 2000, 1791– 1798. (4) Picard, C.; Bosco, M. Soil antimony pollution and plant growth stage affect the biodiversity of auxin-producing bacteria isolated from the rhizosphere of Achillea ageratum L. FEMS Microbiol. Ecol. 2003, 46, 73–80. (5) Murciego, A. M.; Sanchez, A. G.; Gonzalez, M. A. R.; Gil, E. P.; Gordillo, C. T.; Fernandez, J. C.; Triguero, T. B. Antimony distribution and mobility in topsoils and plants (Cytisus striatus, Cistus ladanifer and Dittrichia viscose) from polluted Sb-mining areas in Extremadura (Spain). Environ. Pollut. 2007, 145, 15– 21. (6) Baroni, F.; Boscagli, A.; Protano, G.; Riccobono, F. Antimony accumulation in Achillea ageratum, Plantago lanceolata and Silene vulgaris growing in an old Sb-mining area. Environ. Pollut. 2000, 109, 347–352. (7) Flynn, H. C.; Meharg, A. A.; Bowyer, P. K.; Paton, G. I. Antimony bioavailability in mine soils. Environ. Pollut. 2003, 124, 93–100. (8) Johnson, C. A.; Moench, H.; Wersin, P.; Kugler, P.; Wenger, C. Solubility of antimony and other elements in samples taken from shooting ranges. J. Environ. Qual. 2005, 34, 248–254. (9) Takaoka, M.; Fukutani, S.; Yamamoto, T.; Horiuchi, M.; Satta, N.; Takeda, N.; Oshita, K.; Yoneda, M.; Morisawa, S.; Tanaka, T. Determination of chemical form of antimony in contaminated soil around a smelter using X-ray absorption fine structure. Anal. Sci. 2005, 21, 769–773. (10) Scheinost, A. C.; Rossberg, A.; Vantelon, D.; Xifra, I.; Kretzschmar, R.; Leuz, A. K.; Funke, H.; Johnson, C. A. Quantitative antimony
(19)
(20)
(21)
(22)
(23)
(24)
(25)
(26)
speciation in shooting-range soils by EXAFS spectroscopy. Geochim. Cosmochim. Acta 2006, 70, 3299–3312. Filella, M.; Belzile, N.; Chen, Y. W. Antimony in the environment: a review focused on natural waters II. Relevant solution chemistry. Earth Sci. Rev. 2002, 59, 265–285. Belzile, N.; Chen, Y. W.; Wang, Z. J. Oxidation of antimony (III) by amorphous iron and manganese oxyhydroxides. Chem. Geol. 2001, 174, 379–387. Leuz, A. K.; Monch, H.; Johnson, C. A. Sorption of Sb(III) and Sb(V) to goethite: Influence on Sb(III) oxidation and mobilization. Environ. Sci. Technol. 2006, 40, 7277–7282. Mitsunobu, S.; Harada, T.; Takahashi, Y. Comparison of antimony behavior with that of arsenic under various soil redox conditions. Environ. Sci. Technol. 2006, 40, 7270–7276. Nakamaru, Y.; Tagami, K.; Uchida, S. Antimony mobility in Japanese agricultural soils and the factors affecting antimony sorption behavior. Environ. Pollut. 2006, 141 (2), 321–326. Cornelis, G.; Van Gerven, T.; Vandecasteele, C. Antimony leaching from uncarbonated and carbonated MSWI bottom ash. J. Hazard. Mater. 2006, 137, 1284–1292. Hammel, W.; Steubing, L.; Debus, R. Assessment of the ecotoxic potential of soil contaminants by using a soil-algae test. Ecotoxicol. Environ. Saf. 1998, 40, 173–176. He, M. C.; Yang, J. R. Effects of different forms of antimony on rice during the period of germination and growth and antimony concentration in rice tissue. Sci. Total Environ. 1999, 244, 149– 155. Murata, T.; Kanao-Koshikawa, M.; Takamatsu, T. Effects of Pb, Cu, Sb, in and Ag contamination on the proliferation of soil bacterial colonies, soil dehydrogenase activity, and phospholipid fatty acid profiles of soil microbial communities. Water Air Soil Pollut. 2005, 164, 103–118. Kuperman, R. G.; Checkai, R. T.; Simini, M.; Phillips, C. I.; Speicher, J. A.; Barclift, D. J. Toxicity benchmarks for antimony, barium, and beryllium determined using reproduction endpoints for Folsomia candida, Eisenia fetida, and Enchytraeus crypticus. Environ. Toxicol. Chem. 2006, 25, 754–762. Speir, T. W.; Ketlles, H. A.; Percival, H. J.; Parshotam, A. Is soil acidification the cause of biochemical responses when soils are amended with heavy metal salts? Soil Biol. Biochem. 1999, 31, 1953–1961. Stevens, D. P.; McLaughlin, M. J.; Heinrich, T. Determining toxicity of lead and zinc runoff in soils: salinity effects on metal partitioning and on phytotoxicity. Environ. Toxicol. Chem. 2003, 22, 3017–3024. Smolders, E.; Brans, K.; Coppens, F.; Merckx, R. Potential nitrification rate as a tool for screening toxicity in metalcontaminated soils. Environ. Toxicol. Chem. 2001, 20, 2469– 2474. ISO 11269-2. Soil qualitysDetermination of the effects of pollutants on soil florasPart2: Effects of chemicals on the emergence and growth of higher plants; International Organization for Standardization: Geneva, Switzerland, 1995. ISO 11269-1. Soil qualitysDetermination of the effects of pollutants on soil florasPart 1: Method for the measurement of inhibition of root growth; International Organization for Standardization: Geneva, Switzerland, 1995. Krachler, M.; Emons, H. Urinary Antimony Speciation by HPLCICP-MS. J. Anal. Atom. Spectrom. 2001, 16, 20–25.
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