Mercury accumulation and speciation in plants and ...

Geoderma 253–254 (2015) 30–38

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Mercury accumulation and speciation in plants and soils from abandoned cinnabar mines Rodolfo Fernández-Martínez a, Raquel Larios a,1, Isabel Gómez-Pinilla b, Belén Gómez-Mancebo a, Sol López-Andrés c, Jorge Loredo d, Almudena Ordóñez d, Isabel Rucandio a,⁎ a

Centro de Investigaciones Energéticas, Medioambientales y Tecnológicas (CIEMAT), Av. Complutense, 40, 28040 Madrid, Spain Centro de Asistencia a la Investigación de Técnicas Geológicas, Universidad Complutense de Madrid, C/José Antonio Novais s/n°, 28040 Madrid, Spain Departamento de Cristalografía y Mineralogía, Facultad de Ciencias Geológicas, Universidad Complutense de Madrid, C/José Antonio Novais s/n, 28040 Madrid, Spain d Dpto. de Explotación y Prospección de Minas, Universidad de Oviedo, ETS Ingenieros de Minas, Independencia, 13, E-33004 Oviedo, Spain b c

a r t i c l e

i n f o

Article history: Received 15 December 2014 Received in revised form 1 April 2015 Accepted 6 April 2015 Available online xxxx Keywords: Mercury Cinnabar mines Speciation Soil pollution Plant uptake Transfer factors

a b s t r a c t This study examines total Hg, free Hg(0), matrix-bound Hg and MeHg contents in soils and plant tissues from two old cinnabar mining sites (La Soterraña and Los Rueldos) in Asturias (Spain), as well as Hg transfer and translocation from soils to plants. The studied soils from both mines accumulated moderate to very high total Hg concentrations (36–1709 mg·kg−1) but quite low available Hg contents (0.005–3.062 mg·kg−1) which resulted in relatively low transfer factor values. Matrix-bound Hg was the prevalent Hg form in soils from both mining sites representing 67–88% of total Hg content. Significant Hg(0) concentrations were found in soils as a consequence of atmospheric deposition. Appreciable MeHg concentrations were found in soils from La Soterraña mining site while non-detectable MeHg could be found in soils from Los Rueldos. All the studied plants can be considered as excluders. Hg(0) is practically absent in roots indicating that this Hg form is not uptaken from soils. Hg(0) contents found in aboveground tissues evidenced that foliar uptake of atmospheric Hg occurs. Significant MeHg contents were found in the aerial parts for all studied plants. However, MeHg contents in roots were extremely low in La Soterraña plants and non-detectable in those from Los Rueldos. © 2015 Elsevier B.V. All rights reserved.

1. Introduction Widespread Hg contamination in the environment has occurred due to anthropogenic activities including mining and smelting, among others (ATSDR, 1999; Seiler et al., 2004; Stopford, 1979; USEPA, 1984). Mercury, as well as other trace elements, circulates between different environmental compartments (i.e. atmospheric aerosol, dust, soil, plants, sediment) in gas phase, in aqueous solution, and as particulate solids (Charlesworth et al., 2011). Historic cinnabar mines represent persistent mercury sources to all environmental compartments. Smelting activities release large quantities of Hg to the atmosphere, mostly as Hg(0) vapor, and also secondary Hg compounds can be eventually dumped into the surrounding soils where can they be effectively complexed by soil components as humic acids and iron oxyhydroxides (Biester et al., 1999; Kocman et al., 2011; Navarro et al., 2006). Furthermore, natural weathering of waste materials stored in spoil heaps can mobilize Hg to the surrounding soils, waters and biota (Loredo et al., 2006). Mercury concentrations in soils usually range between 0.01

⁎ Corresponding author. E-mail address: [email protected] (I. Rucandio). 1 Current address: LGC Limited, Queens Road, Teddington, Middlesex TW11 0LY, UK.

http://dx.doi.org/10.1016/j.geoderma.2015.04.005 0016-7061/© 2015 Elsevier B.V. All rights reserved.

and 0.2 mg·kg−1 (Adriano, 2001), but these are significantly higher in soils affected by Hg mining. Many studies have demonstrated that mercury can be uptaken by plants by assimilation from topsoil into roots via transpiration stream. Hg availability for plants is usually low because of its low solubility in the soil solution (Baya and Van Heyst, 2010), and it is mainly accumulated in roots (Patra and Sharma, 2000). However, the transfer and translocation of Hg(II) and the highly toxic MeHg species from soils to shoots can occur (Bishop et al., 1998). Plants can also uptake Hg from the atmosphere which may occur by direct absorption through the leaves, by means of stomata, or foliar adsorption of wet and dry deposited Hg(0) (Millhollen et al., 2006). Then, Hg(0) vapor can be reemitted to the atmosphere from leaves (Fay and Gustin, 2007). Dry deposition of ionic mercury compounds, including the highly toxic MeHg species, and particulate bound mercury is also possible (Risch et al., 2012; Witt et al., 2009). Therefore, plants may act as a significant pathway by which mercury may enter or leave terrestrial ecosystems (Mohamed et al., 2003). Since elevated Hg concentrations are found in soils and atmosphere at historic mining sites, Hg accumulation in plants growing in such areas can be unusually high (Higueras et al., 2003, 2006; Kocman et al., 2011). Hence, the evaluation of Hg accumulation and its speciation in plants from historic mining areas is of great concern. Mercury behavior in

R. Fernández-Martínez et al. / Geoderma 253–254 (2015) 30–38

the soil–plant systems can be studied by means of soil-to-plant transfer factors. The transfer factors depend not only on the plant species but also on the Hg concentration, its speciation and its availability. It has been demonstrated that non-cinnabar Hg forms are more available to living organisms as earthworms (Han et al., 2012). Geology, mineralogy, fractionation of soils and sediments and a detailed monitoring of soils and waters have been widely investigated in the study area (Fernández-Martínez et al., 2014; Loredo et al., 2005, 2010; Ordonez et al., 2011). Nevertheless, few studies have been carried out about the impact of Hg in vegetation (Ordóñez et al., 2013) and the Hg assimilation and speciation in plants. Then, the main objectives of this work were: (i) to evaluate the degree of accumulation of total Hg in soils and plants grown in two abandoned cinnabar mines of Asturias (Northern Spain), (ii) to determine free Hg(0), matrix-bound Hg and MeHg concentrations in soils and differentiated parts of several plant species in order to estimate the Hg accumulation due to soilto-plant transfer and foliar deposition, and (iii) to assess Hg availability, transfer and translocation from soils to plants by means of transfer factors.

2. Materials and methods 2.1. Study site Mercury mining was an important activity in Asturias (Northern Spain), to the point of becoming the third world Hg producer during the 1950–70s (Loredo et al., 1988). An intense mining and metallurgical activity was intermittently developed in the region from Roman times until the 1970s, when all the Asturian Hg mines were abandoned, in the absence of a specific environmental legislation. The legacy of the Hg mining activities remains now in the form of old industrial installations (shafts, mine buildings, roasting furnaces, chimneys, etc.) and mining and metallurgical wastes piled in spoil heaps that constitute a potential source for the spread of Hg to the different environmental compartments. In the studied area, extremely high Hg concentrations have been found in spoil heap wastes (up to 3260 mg·kg−1), soils (up to 700 mg·kg−1) and local stream sediments (up to 3270 mg·kg−1). In contrast, waters are not highly polluted by Hg (b 0.1 μg·L−1 in surface water) which shows that Hg leaching does not occur to a high extent; notwithstanding, concentrations up to 14 μg·L−1 have been found in groundwater sampled in boreholes drilled on the mining site, although Hg was not detected in springs downstream of it. Regarding atmospheric Hg, values up to 2.3 μg·Nm−3, which are 10 times higher than the local background, were found in this area (Loredo et al., 2007; Ordóñez et al., 2013). For this work, plants from two abandoned Asturian Hg mine sites, La Soterraña (Pola de Lena district) and Los Rueldos (Mieres district), were collected and studied (Fig. 1). La Soterraña Hg mine was the second most important one in Asturias. The mineralogy of ore deposit includes cinnabar, realgar and pararealgar, orpiment, As-rich pyrite, marcasite and arsenopyrite as primary minerals (Loredo et al., 2006). The extracted ore was crushed and roasted on site. Old industrial installations and a big spoil heap (covering 17,000 m2) with significant quantities of wastes from mining and metallurgical operations, still remain on site, exposed to weathering and consequent mobilization of pollutants (Loredo et al., 2006). Los Rueldos mine was not so relevant and metallurgy was not carried out on site. Mineralogy of the ore deposit includes cinnabar, pyrite (occasionally As-rich pyrite), melnikovite, sphalerite, marcasite, chalcopyrite, arsenopyrite, galena, stibnite and realgar as primary minerals. A spoil heap of mine wastes covering an area of 1340 m2 remains on site (Loredo et al., 2005). A low flow of acid mine drainage (AMD) arises from a mine gallery and it forms a small pond. This AMD, with pH around 2.5, contains up to 4600, 9.2, 0.5 and 0.014 mg·L−1 of sulfate, As, Pb and Hg, respectively (Loredo et al., 2005).

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2.2. Plant and soil sampling and preparation The varied flora around La Soterraña site includes oligotrophic forests with oak and birch, conifer plantations (Monterrey pine-tree) and hardwood plantations (chestnut tree), as well as silicicolous fern, heather, meadows and pastures. The same shrubby vegetation is found near Los Rueldos too, also with predominance of meadows, along with chestnut trees (Asturias, 2001). Right on mining the sites, sub-shrubby species, such as fern and herbaceous perennial plants are predominant. The most abundant plant species at the studied sites were collected, as well as the soils where they were growing. In La Soterraña two points were sampled: P1 and P3, located above and below the spoil heap, respectively (Fig. 1). Crupina vulgaris Cass was the plant species collected at P1 (sample P1-E1). Three plant species were sampled in P3: Typha latifolia, Phyllitis scolopendrium and Dryopteris filix-mas (samples P3E4, P3-E5 and P3-H6, respectively). In Los Rueldos, two plant species were collected at point P8, close to the AMD pond (Fig. 1). These were Calluna vulgaris (L.) Hull (sample P8-E7) and Dryopteris affinis (sample P8-H7). Once collected, each plant sample was vigorously washed with tap water and rinsed with deionized water in the laboratory. They were divided in aboveground mass and roots. They were airdried for 7 days to ensure constant weight and afterward ground by using a laboratory grinder. The soils where the sampled plants had grown were collected at a depth of 25 cm when possible, and stored in polyethylene bags. Samples were disaggregated in the laboratory and the non-mineral material and rock fragments higher than 2 mm were discarded. Then, they were oven dried at a temperature below 40 °C to minimize loss of volatile elements, such as Hg during 7–15 days. Afterwards, a split sample for study by XRD was crushed in an agate mortar and sieved to a size finer than 53 μm (Zhao et al., 2007). 2.3. Analytical methods The elemental composition of the sampled soils was determined by wavelength dispersive X-ray fluorescence spectrometry using an AXIOS automated X-Ray Fluorescence (XRF) spectrometer (PANalytical) equipped with a 4 kW Rh tube. Sieved sub-samples were pressed into thick pellets of 37 mm diameter using “elvacite” (C8H14O2) as a binder. The certified reference materials (CRM) NIST 2710 Montana soil and IAEA 405 were analyzed following the same procedure. The obtained results differed less than 10% in the case of the major elements and less than 5% for minor elements compared to the certified values. The X-Ray Diffraction (XRD) patterns were obtained using a Bruker D8 ADVANCE diffractometer with Cu-Kα (λ = 1.54 Å) radiation. To determine the bulk mineralogy, randomly oriented powder of every sample was analyzed from 2 to 65° 2Θ at a speed of 0.02° 2Θ/1 s. A current of 30 mA and a voltage of 40 kV were employed as tube setting. The mineralogical composition and relative amount of every phase were determined following the procedure of Chung, (1974a,b). XRD analyses of oriented samples were performed in order to estimate the overall mineralogical composition of the clay fraction (b2 μm). The fine fraction was extracted from the soil samples after destruction of organic matter by using H2O2 at room temperature. Carbonates were removed by adding the CH3COOH–CH3COONa buffer (pH 5) and washing with distilled water. Clay dispersion was accomplished with mechanical treatment. The b2 μm clay fraction was separated by sedimentation procedure on soil samples previously ultrasonically dispersed in ultrapure water. Oriented specimens were prepared by depositing a clay suspension onto a glass slide. The samples were scanned from 2 to 35° 2Θ at a speed of 0.02° 2Θ/1 s after air-drying, solvation with ethylene glycol at 60 °C for 48 h and heating at 550 °C for 2 h according to Robert and Tessier Robert and Tessier (1974). pH and redox potential (Eh) of soils were measured with a Pt–Ag/ AgCl selective electrode (CRISON) on sample/water suspensions at 1:2.5 ratio (w/v) (McLean, 1982). Total Organic Content (TOC) was

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Spoil heap

P1 P8 Spoil heap

P3

Fig. 1. Location of sampling sites.

determined as the difference between total carbon (TC) and Total Inorganic Carbon (TIC). A LECO CS-244 elemental analyzer was used for the TC determination in the soil samples. The carbon content of samples was determined by combustion in an induction furnace in the presence of oxygen gas. The amount of carbon dioxide was measured by an infrared detection method. TIC was determined using a Shimadzu-VCHS instrument equipped with a solid sample combustion unit. Samples were acidified with phosphoric acid and sparged with zero air to convert inorganic carbon to volatilized carbon dioxide. The CO2 formed was transported to the detector in a carrier gas stream and measured directly by an infrared detector. The measured CO2 consists of carbon derived from carbonates, hydrogen carbonates and dissolved carbon dioxide in the sample. 2.3.1. Total Hg, free elemental Hg, matrix-bound Hg and available Hg Total Hg determinations in soils and plant tissue samples were performed by using a DMA-80 instrument (Milestone, Sorisole, Italy) following the recommendations given by the EPA 7473 method (EPA, 2007). To ensure the quality of the results, the same procedure was applied to the CRM NIST 2710 Montana soil (highly elevated trace element concentrations) and NIST 1575 Pine needles with total Hg contents of 32.6 ± 1.8 mg·kg−1 and 0.15 ± 0.05 mg·kg−1, respectively. The obtained results were 30.9 ± 0.8 mg·kg−1 and 0.13 ± 0.02 mg·kg−1, in good agreement with the certified values. Free elemental Hg (free Hg(0)) and matrix-bound Hg in soil and plant samples were sequentially evaluated by means of a pyrolysis

technique by following procedures reported in the literature (Navarro et al., 2006; Sladek and Gustin, 2003). 1 g sample aliquots were placed in porcelain crucibles and first heated at 80 °C for thermal desorption of free elemental Hg according to the recommendations of Sladek and Gustin (2003). A portion of sample was taken to Hg determination and free elemental Hg(0) was calculated from the difference between total Hg contents before and after the heating process. The rest of sample was then heated at 250 °C during 48 h for thermal desorption of matrix-bound Hg (Biester and Scholz, 1997; Navarro et al., 2006). Afterwards, samples were analyzed for determination of the remaining Hg content. Matrix-bound Hg was calculated from the difference between total Hg contents before and after the second heating process. Hg availability in soils was evaluated following the procedure recommended by Jing et al. (2008). A 0.1 M HCl solution was used as an extracting reagent for a soil/solution of 1:5 ratio. 5.0 g of soil was weighed into a centrifuge tube and then 25 mL of ultrapure water was added. Samples were stirred for 30 min by end-over-end rotation and then centrifuged at 5000 rpm for 15 min. Supernatants were removed, filtered through a 0.45 μm nylon syringe filter and Hg determined by the DMA-80 instrument. 2.3.2. MeHg extraction in soils and plant tissues The extraction procedure was adapted from a previous one developed by the authors (Fernández-Martínez and Rucandio, 2013). 0.5 g of sample was weighed into a 40 mL polyethylene centrifuge tube and 10 mL of 0.3 M of CuBr2 solution in 50% v/v HCl was added. The mixture


was mechanically shaken by an end-over-end apparatus for 15 min. Then 5 mL of toluene was added and the sample was shaken for another 15 min. After centrifuging at 8000 rpm for 15 min, 3 mL of the organic phase was collected and transferred to a 40 mL polyethylene centrifuge tube. 5 mL of toluene was newly added to the sample and it was shaken and centrifuged again. 5 mL of organic phase was collected and combined with the previous supernatant. 5 mL of 3.7 mM N-acetyl-Lcysteine (NAC) solution was added to the centrifuge tube with the toluene extracts for back extraction and then shaken for 15 min. After centrifugation at 8000 rpm for 15 min, 2 mL of NAC phase was collected using a HPLC syringe, filtered by 0.22 μm syringe filters and stored in glass vials until analysis. 2.3.3. Mercury speciation Separation and determination of MeHg in the extracts was carried out by coupled high performance liquid chromatography–on-line ultraviolet assisted oxidation–cold vapor-atomic fluorescence spectrometry (HPLC–UV–CV-AFS). The HPLC system consisted on a Varian Prostar ternary solvent delivery module model 230 equipped with a Rheodyne injector model 7725 and a 200 μL loop for sample injection. The instrument used in the combination of vapor generation and atomic fluorescence spectrometry was a PSA Millennium Merlin from PS Analytical. The chromatographic separation of Hg(II) and MeHg was achieved using a C-18 reversed-phase column (Hypersil ODS-2 4.6 × 250 mm, 5 μm particle size). A 10 mM NAC in 0.06 M Ammonium Acetate solution buffered at pH 5.5 was used as the mobile phase at a flow rate of 0.7 mL·min−1. The separated Hg species were subjected to UV assisted post-column oxidation by using a 0.002 M KBrO3 + 0.002 M KBr solution at 1.9 mL·min−1 flow rate and Hg vapor generation with 2% w/v SnCl2 in 10% v/v HCl solution at 1.9 mL·min−1 flow rate. The gas–liquid mixture was delivered into a gas–liquid separator, and gas compounds were swept from the mixture with a flow of argon, dried and fluorescence emission detected at 254 nm. Detection limits, calculated as three times the standard deviation of a blank injected ten times divided by the sensitivity, were 0.5 μg·kg−1 for MeHg. The suitability of the method and the accuracy of the obtained results were checked by analyzing two certified reference materials (CRM) BCR-580 estuarine sediment and NIST 1946 Lake Superior Fish Tissue with certified MeHg contents of 75.5 ± 3.7 μg·kg−1 and 394 ± 15 μg·kg− 1, respectively. The obtained results were 72.8 ± 2.8 and 379 ± 11 μg·kg−1 respectively. MeHg recoveries higher than 95% of certified value were obtained for both CRMs. Then, the methodology can be considered suitable for Hg speciation in geological and biological samples. 3. Results and discussion 3.1. Soils characterization Multielemental analysis in soil samples (Table 1) shows high concentrations in elements not characteristic of natural soil, such as Ce, Ga, La, Nd, Rb, Sr, Th, and U, which reveals the influence of anthropogenic factors on the enrichment of these elements. Given the proximity of the samples to the spoil heaps, they are actually a mixture of soil and leached waste that have been integrated for ages, particularly in Los Rueldos. This behavior agrees with that found in samples from other abandoned mine sites in Asturias (Fernández-Martínez et al., 2006). As, Nd, Pb, Sb and W contents are significantly higher in sample P8 (Los Rueldos) than in samples P1 and P3 (La Soterraña), whereas the elements Ca, F, Co, Ge, Hg, Mn and Zn show the opposite trend. In particular, the high Pb content of sample P8-H7 must be related to the presence of galena or Pb secondary minerals such as cerussite in the soilwaste mixture. Mn concentrations were significantly higher in samples P3. Some elements (As, Hg and also Sb) with analogous geochemical behavior were found at a high level of concentration in some samples. Regarding the chemical properties of soils (Table 1), samples from La Soterraña exhibited weakly alkaline pHs, typical from the calcareous

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soils in this area (Loredo et al., 2006), which is in agreement with the appreciable amounts of calcium found in these samples. Redox potentials of these sediments were generally weakly oxidant but slightly reducing in some samples at La Soterraña site. TOC contents in this area ranged between 3.6 and 21.8%. However, the situation in Los Rueldos was completely different. The acidic pHs, high Ehs and low TOC contents in soils indicate that AMD occurs in this site. Sulfur contents – indicative of the presence of pyrite and other sulfides – were higher than those found in La Soterraña soils. It is noteworthy the very low calcium contents in these samples, suggesting the non-calcareous nature of soils and rocks at this site. According to the X-ray diffraction analysis data (Fig. 2), the samples showed abundant quartz and phyllosilicate contents. The presence of calcite was detected in a greater or lesser extent in La Soterraña samples and peaks of fluorite appeared in P3-E4 and P3-E5 (La Soterraña). Additional reflections were observed corresponding to some minor minerals such as cerussite, and arsenic and antimony oxides, mainly in P8-E7 sample. Phyllosilicates are distributed in kaolinite and mica-illite as major compounds and occasionally vermiculite. The important variability of the results of both XRD and XRF in the samples collected at the same sampling point could be attributed to the heterogeneity of the soils. In both areas, the collected soils were a mixture of developed soil and mineral residues of the ore activities. Similar results at both locations were also observed by other authors (Ordóñez et al., 2013). 3.2. Total mercury 3.2.1. Soils Total Hg concentrations in plant tissues and its corresponding soils are summarized in Table 2. The soil samples from both studied areas displayed a wide range of Hg concentrations, from 36 to 1709 mg·kg− 1. These concentrations were similar to those found in other historic mining districts around the world such as Almadén (Spain), where Hg soil concentrations were found ranging between 0.13 and 2695 mg·kg−1 (Molina et al., 2006); Idrija (Slovenia), where Hg in alluvial soils range between 0.595 and 1970 mg·kg− 1 (Gosar and Zibret, 2011) or Wanshan (China) with total Hg contents in soils ranging between 5.1 and 790 mg·kg−1 (Horvat et al., 2003). Hg values in all sampling points were largely higher than those reported in literature to consider the soils as toxic (0.3–5 mg·kg−1) (Kabata-Pendias and Pendias, 1992; Poschenrieder and Barceló, 2003). 3.2.2. Plants Hg contents in the studied plants ranged between 0.6 and 48 mg·kg−1 in aerial parts and from 1.8 to 160 mg·kg− 1 in roots (Table 2). The plant species with the highest amount of total Hg was C. vulgaris from La Soterraña site, with 160 mg·kg− 1 in roots and 48 mg·kg−1 in the aboveground mass. This high concentration is comparable to that found in other herbaceous plant species (Marrubium vulgare L.) in other historic Spanish cinnabar districts (Garcia-Sanchez et al., 2009; Higueras et al., 2003). Cuprina vulgaris was collected in the most polluted soil, which seems to indicate that the extent of mercury accumulation depends directly on the degree of pollution of the site where they are growing. However, for the rest of the plants, despite the fact that they grow in highly polluted soils, they did not accumulate so much mercury compared to C. vulgaris. This suggests that mercury accumulation depends not only on the soil pollution, but also on the considered plants and its ability to incorporate it. As for soils, most of collected plant samples exceeded the Hg critical concentration in plants, that is, the level above which toxicity effects are likely which is 1–3 mg·kg− 1(Kabata-Pendias and Pendias, 1992). The symptoms of mercury toxicity depends on each plant species, but commonly are the inhibition of photosynthesis, stunted roots, and stunted seedlings, which can provoke reductions in yield. The accumulation of mercury

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Table 1 Multielemental analysis, pH, Eh and TOC results from studied soils. La Soterraña

Los Rueldos

Elements

Units

P1-E1

P3-E4

P3-E5

P3-H6

P8-E7

P8-H7

Al As C Ca Cl F Fe K Mg Na P S Si Ti Ba Br Ce Co Cr Cs Cu Ga Ge Hg I La Mn Mo Nb Nd Ni Pb Rb Sb Sc Se Sm Sr Th U V W Y Yb Zn Zr pH Eh TOC

% % % % % % % % % % % % % % mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1 mg·kg−1

9.4 0.67 3.6 2.1 0.024 0.44 4.4 1.6 b0.005 0.14 0.080 0.15 25 0.45 495 6.6 113 20 217 8.4 40 7.9 8.3 1709 11 50 590 3.3 16 48 59 52 99 51 17 3.2 4.6 259 13 11 156 3.8 26 4.2 124 190 7.64 −26 3.3

1.6 0.050 22 1.1 0.012 0.46 0.41 0.3 b0.005 0.057 0.084 0.11 4.5 0.045 261 174 42 11 100 8.3 32 7.5 5.5 91 82 23 1461 1.8 9 19 70 25 45 34 12 3.0 1.9 332 4.6 3.3 112 b2 16 b4 152 102 7.68 −90 21.2

6.3 0.95 8.2 5.3 0.016 1.7 2.9 1.1 0.38 0.15 0.123 0.22 18 0.28 447 35 127 19 446 3.8 41 4.3 4.3 128 21 39 1089 3.0 13 44 90 41 62 77 20 6.3 7.5 375 8.8 12 134 4.6 22 b4 162 167 7.77 6 7.1

9.6 0.10 4.2 0.96 b0.005 0.22 4.4 2.0 0.52 0.31 0.071 0.13 25 0.47 412 8.2 85 19 214 7.7 29 17 1.1 41 b8 45 978 2.5 17 37 63 32 99 6 15 0.7 7.3 169 13 3.7 134 3.6 28 b4 85 256 7.67 32 4.2

9.9 2.1 1.7 0.016 0.007 b0.005 3.1 1.4 0.45 0.058 0.064 0.42 28 0.47 334 b1.1 83 4.1 303 1.9 25 15 b0.8 146 b8 55 52 2.2 16 72 51 1975 54 1085 28 10 8.3 172 9.0 3.8 159 10 29 6.4 22 263 3.85 228 1.7

11 1.1 1.6 0.030 0.011 b0.005 5.2 2.2 0.37 0.11 0.073 0.76 26 0.50 390 1.8 110 7.2 140 5.2 16 25 b0.8 36 b8 61 127 2.0 18 58 25 95 91 44 11 5.5 6.3 196 14 2.9 205 10 25 4.6 19 156 3.87 285 0.3

mV %

in roots seems to inhibit the uptake of other elements, such as potassium (Kabata-Pendias and Pendias, 1992). 3.3. Free Hg(0), matrix-bound Hg and available Hg in soils Free Hg(0) contents in soils ranged between 2.1 and 113 mg·kg−1 (Table 2), with percentages varying between 6 and 20% of the respective total Hg contents. The presence of free Hg(0) indicated atmospheric Hg deposition in the studied areas (Guedron et al., 2009). It has been reported that calcines and mine wastes located in Hg mining activities areas can be major sources of atmospheric Hg, in the form of Hg(0), which can be subsequently deposited in surrounding soils (Dai et al., 2012). On the other hand, matrix-bound Hg contents ranged from 27 to 1496 mg·kg−1 with percentages varying between 67 and 88% of total Hg (Table 2). This means that matrix-bound Hg is the most important Hg-bearing phase present in the studied soils. This phase accounts for Hg complexed to soil components with the exception of the highly stable cinnabar species, which is typically released at temperatures above

280 °C (Biester and Scholz, 1997). The prevalence of this phase is not surprising since similar behaviors have been observed in other cinnabar districts (Kocman et al., 2004; Navarro et al., 2006). According to similar studies matrix-bound Hg can be attributed mainly to metallic Hg, formed during roasting process and adsorbed to mineral matrix components such as carbonates, Fe oxyhydroxides and humic acids. In addition, metallic Hg associated to humic acids represents an environmental risk since it can be further methylated or oxidized to reactive Hg(II) (Gray et al., 2004). Hg availability was very low in soils from both mining areas, representing only between 0.001 and 0.66% of total Hg content (Table 2). These results are comparable to those reported from Almadén (0.005–0.03%) or Usagre mining districts (0.005–0.200%) (Garcia-Sanchez et al., 2009; Moreno-Jimenez et al., 2006). Higher available Hg concentrations were found in soils from La Soterraña, which is in agreement with the higher Hg concentrations found in the roots of the plants collected in this area. Hg was not easily leachable in soils from Los Rueldos, despite being heavily affected by AMD (Table 2). The refractory nature of Hg in these mining districts might be expected, as previous studies based on the application of sequential


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Mineralogical patterns and qualitative composition of the bulk sample Chemical Composition Qtz Quartz

SiO2

Cal Calcite

CaCO3

Flu Fluorite

CaF2

Cer Cerussite

PbCO3

P1-E1-S P3-E4-S P3-E5-S P3-H6-S P8-E7-S P8-H7-S

Phy Phylosilicate Ars Arsenolite

As2O3

Mineralogical patterns and semiquantitative composition of the clayf raction (