Arsenic and antimony contamination of waters, stream sediments and soils in the vicinity of abandoned antimony mines in the Western Carpathians, Slovakia

Applied Geochemistry 27 (2012) 598–614

Contents lists available at SciVerse ScienceDirect

Applied Geochemistry journal homepage: www.elsevier.com/locate/apgeochem

Arsenic and antimony contamination of waters, stream sediments and soils in the vicinity of abandoned antimony mines in the Western Carpathians, Slovakia Edgar Hiller a,⇑, Bronislava Lalinská b, Martin Chovan b, Lˇubomír Jurkovicˇ a, Tomáš Klimko b, Michal Jankulár a, Róbert Hovoricˇ b, Peter Šottník c, Renáta Flˇaková d, Zlatica Zˇenišová d, Ivana Ondrejková d a

Department of Geochemistry, Comenius University, Mlynská dolina, SK-842 15 Bratislava, Slovak Republic Department of Mineralogy and Petrology, Comenius University, Mlynská dolina, SK-842 15 Bratislava, Slovak Republic Department of Geology of Mineral Deposits, Comenius University, Mlynská dolina, SK-842 15 Bratislava, Slovak Republic d Department of Hydrogeology, Comenius University, Mlynská dolina, SK-842 15 Bratislava, Slovak Republic b c

a r t i c l e

i n f o

Article history: Received 4 February 2011 Accepted 10 December 2011 Available online 16 December 2011 Editorial handling by E.L. Ander

a b s t r a c t Environmental contamination with As and Sb caused by past mining activities at Sb mines is a significant problem in Slovakia. This study is focused on the environmental effects of the 5 abandoned Sb mines on water, stream sediment and soil since the mines are situated in the close vicinity of residential areas. Samples of mine wastes, various types of waters, stream sediments, soils, and leachates of the mine wastes, stream sediments and selected soils were analyzed for As and Sb to evaluate their geochemical dispersion from the mines. Mine wastes collected at the mine sites contained up to 5166 mg/kg As and 9861 mg/kg Sb. Arsenic in mine wastes was associated mostly with Fe oxides, whereas Sb was present frequently in the form of individual Sb, Sb(Fe) and Fe(Sb) oxides. Waters of different types such as groundwater, surface waters and mine waters, all contained elevated concentrations of As and Sb, reaching up to 2150 lg/L As and 9300 lg/L Sb, and had circum-neutral pH values because of the buffering capacity of abundant Ca- and Mg-carbonates. The concentrations of Sb in several household wells are a cause for concern, exceeding the Sb drinking water limit of 5 lg/L by as much as 25 times. Some attenuation of the As and Sb concentrations in mine and impoundment waters was expected because of the deposition of metalloids onto hydrous ferric oxides built up below adit entrances and impoundment discharges. These HFOs contained >20 wt.% As and 1.5 wt.% Sb. Stream sediments and soils have also been contaminated by As and Sb with the peak concentrations generally found near open adits and mine wastes. In addition to the discharged waters from open adits, the significant source of As and Sb contamination are waste-rock dumps and tailings impoundments. Leachates from mine wastes contained as much as 8400 lg/L As and 4060 lg/L Sb, suggesting that the mine wastes would have a great potential to contaminate the downstream environment. Moreover, the results of water leaching tests showed that Sb was released from the solids more efficiently than As under oxidizing conditions. This might partly explain the predominance of Sb over As in most water samples. Ó 2011 Elsevier Ltd. All rights reserved.

1. Introduction Antimony and As are metalloids widely distributed in the environment. Both Sb and As are considered to be highly toxic for biota and humans, even when they are present at low concentrations in the environment (Gebel, 1997, 1999; Filella et al., 2002; Smedley and Kinniburgh, 2002; Reimann et al., 2009). Compounds of Sb and As have been declared as pollutants of priority interest by the European Community Directive 76/464/EEC for three decades (Council of the European Communities, 1976). Concentrations of

⇑ Corresponding author. Tel.: +421 02 60296 218; fax: +421 2 602 96 217. E-mail address: [email protected] (E. Hiller). 0883-2927/$ - see front matter Ó 2011 Elsevier Ltd. All rights reserved. doi:10.1016/j.apgeochem.2011.12.005

As in drinking water, surface waters, stream sediments, and soils are strictly limited by the current EU legislation, but to date only the limit for Sb in drinking water has been established. Elevated concentrations of Sb and As in natural environments have recently received a great scientific and public attention worldwide. Both metalloids commonly occur in the same geochemical environment, although the geochemical behavior of As differs significantly in many aspects from that of Sb (Belzile et al., 2001; Tighe et al., 2005; Casiot et al., 2007). Exploitation and processing of hydrothermal Sb ores are believed to be one of the most important sources for environmentally significant Sb and As contamination (Baroni et al., 2000; Ashley et al., 2003, 2006; Wilson et al., 2004a). There are many abandoned mines worldwide where no cleanup of wastes has occurred. Thus,

E. Hiller et al. / Applied Geochemistry 27 (2012) 598–614

they represent sites of serious contamination associated with the release of As and Sb mainly through oxidative weathering of exposed sulfide minerals (Craw et al., 2004; Wilson et al., 2004b; Alvarez et al., 2006; Haffert and Craw, 2008a; Filella et al., 2009). Substantial Sb deposits and smaller occurrences are located in various zones of the Western Carpathians of Slovakia (Chovan et al.,

599

1994). Generally, the hydrothermal activity produced ores with Au-bearing pyrite (FeS2) and arsenopyrite (FeAsS) and Sb ores with dominant stibnite (Sb2S3) (Chovan et al., 1992). All Slovak Sb deposits were closed and abandoned in the early 1990s without remediation. Long-term exploitation of Sb deposits in Slovakia has also produced large amounts of waste rocks and tailings which

Fig. 1. Schematic map showing locations of the study areas.

Fig. 2. Map showing the Pernek mine site and locations of sampling points for waters, stream sediments, soils, hydrous ferric oxides and mine wastes.

600

E. Hiller et al. / Applied Geochemistry 27 (2012) 598–614

contribute to the metalloid contamination of the adjacent environment (Veselsky´ et al., 1996; Trtíková et al., 1999; Flˇaková et al., 2005; Majzlan et al., 2007; Jankulár et al., 2008). This study presents the results of geochemical investigations within a multidisciplinary project whose goal was to evaluate the environmental impact of open adits, unconfined waste-rock dumps and tailings left from former mining activities associated with the ˇ ucˇma Sb deposits (ChoPernek, Dúbrava, Medzibrod, Poprocˇ, and C van et al., 2010). Within this subproject, the focus was especially on the chemical composition of waters draining the mining areas, stream sediments, and soils. Laboratory leaching tests were performed in order to estimate the release of As and Sb from selected samples of mined materials (waste rocks), processing residues (tailings), stream sediments, and contaminated soils. This was combined with a mineralogical and geochemical characterization of the different materials also including samples of hydrous ferric oxides (HFOs) commonly precipitating from the adit and mine waste discharges.

2. Description of the studied sites Fig. 1 shows locations of the studied mine sites. The Pernek deposit is situated about 3 km NE of the village of Pernek. The rock complex represents a Devonian volcano-sedimentary formation, later metamorphosed to the amphibolite facies (Cambel and Khun, 1983). The most abundant Sb mineral is stibnite in association with carbonates and quartz, followed by gudmundite, berthierite, pyrite, and arsenopyrite (Kodeˇra et al., 1990; Chovan et al., 1992). Several waste-rock dumps remained around the Pernek deposit. The mine site is drained by the Kostolny´ Creek and its tributary (Fig. 2). The Dúbrava deposit is located in the northern part of the Nízke Tatry Mts, covering both sides of the Krizˇianka Valley, about 7–10 km south of the village of Dúbrava. It was the most significant producer of Sb in the former Czechoslovakia (Chovan et al., 1998). The most abundant ore minerals are stibnite and pyrite. Gangue minerals comprise mostly quartz, calcite, Fe-dolomite, and barite (Michálek and Chovan, 1998). A flotation processing plant was

Fig. 3. Map showing the Dúbrava mine site and locations of sampling points for waters, stream sediments, soils, hydrous ferric oxides and mine wastes.

E. Hiller et al. / Applied Geochemistry 27 (2012) 598–614

operating since 1943 and the flotation waste was deposited in three tailings impoundments (Fig. 3). The mine site is drained by the Paludzˇanka Creek. The Medzibrod deposit is located in the Mocˇiar Valley in the southern part of the Nízke Tatry Mts about 4 km SE of the village of Medzibrod. The ore minerals are mainly stibnite, berthierite, arsenopyrite, zinkenite, chalcopyrite, tetrahedrite, gudmundite, pyrite and galena. The most common non-metallic minerals identified are quartz, carbonates (calcite, dolomite, magnesite and siderite), tourmaline, feldspars and muscovite (Lalinská and Chovan, 2006). The Sb ores were processed by flotation and the waste was deposited in a tailings impoundment (Fig. 4). The Borovsky´ Creek flows through the mineralized area. The Poprocˇ deposit is situated in the southeastern part of the Spišsko-gemerské Rudohorie Mts, in the Petrová Valley, about

601

1 km north of the village of Poprocˇ. Stibnite is the most abundant ore mineral, but pyrite, arsenopyrite and various Pb–Sb–Zn–Cu sulfides occur commonly. The main gangue minerals are quartz, albite, chlorites, sericite, tourmaline, and carbonates, principally Fe-dolomite (Klimko et al., 2009). Tailings from flotation of the sulfide ores were disposed of three impoundments (Fig. 5). The mine site is drained by the Olšava Creek directly receiving mine waters from the abandoned adits and tailing impoundments. The Cˇucˇma deposit is situated in the southern part of the Spišsko-gemerské Rudohorie Mts (eastern Slovakia), near the town of Rozˇnˇava. Primary mineralization is composed of prevalent stibnite, followed by pyrite, arsenopyrite and chalcopyrite. Non-metallic minerals are mainly quartz, carbonates (calcite, siderite and Fedolomite), albite, and tourmaline (Klimko et al., 2009). Tailings produced by the flotation plant were disposed in an impoundment

Fig. 4. Map showing the Medzibrod mine site and locations of sampling points for waters, stream sediments, soils, hydrous ferric oxides and mine wastes.

602

E. Hiller et al. / Applied Geochemistry 27 (2012) 598–614

a WTW pH 340i instrument with SenTixÒORP electrode, respectively. The measured Eh values were corrected against the standard hydrogen electrode (Pitter, 1999). Alkalinity was also measured in the field. Major anions (SO4, NO3 and Cl) were determined by ion chromatography (IC Dionex–120). Metals (Na, K, Ca, Mg, Fe, Mn and Al) were determined by inductively coupled plasma-atomic emission spectrometry (ICP-AES, Varian Liberty 200, Vista AX). Arsenic and Sb concentrations in the water samples were determined by atomic absorption spectrometry (AAS, Varian Spectr AA 220) equipped with a hydride generation system (VGA 76). Instrumental accuracy and precision were monitored using blind duplicates (n = 6) and found to be within the range of ±10% or better. The detection limit was 1.0 lg/L for both As and Sb in waters. 3.2. Solids

Fig. 5. Map showing the Poprocˇ mine site and locations of sampling points for waters, stream sediments, soils, hydrous ferric oxides and mine wastes.

(Fig. 6). The area is drained by two streams called Laz and ˇ ucˇmiansky Creeks, which join together at the upper end of the vilC lage of Cˇucˇma. Essential features of the five hydrothermal Sb deposits studied in this paper are summarized in Table 1.

3. Materials and methods 3.1. Waters The sampling points for water and solid samples are shown in Figs. 2–6. At each mine site, waters were sampled from the discharges of waste-rock dumps, impoundments, and abandoned adits as well as the creeks draining the mine sites. At Dúbrava, Poprocˇ and Cˇucˇma mine sites, samples were also collected from shallow boreholes localized either in tailings impoundments or alluvial sediments below the impoundments. Some samples were also taken from household wells situated downstream of the mine workings. The water samples considered in this study were collected in a single sampling campaign – in June–July 2008. Samples were filtered in situ through a 1 lm pore-size paper filter and then in the laboratory through a 0.45 lm membrane filter. Upon filtration, one aliquot of the waters was acidified with HNO3 for metal analyses and the other one was left unacidified for anion analyses. Temperature and EC, pH and Eh were measured in the field at the time of sampling using a WTW Multi 350i instrument equipped with TetraConÒ325 electrode, SenTixÒ41 electrode, and

Solid grab samples were collected at each mine site and included: (i) mine wastes, (ii) stream sediments, (iii) soils, and (iv) hydrous ferric oxides. For soil samples, 2–3 kg were collected with a plastic scoop from a depth of 0–20 cm. At each mine site, a 3 kg stream sediment sample was taken at 10 cm depth by integrating material from 5 to 10 spots within an area of 1 m2. All HFOs were collected using a plastic scoop and the samples (3–5 L) were stored in plastic bags. Prior to the analyses, the mine waste, stream sediments, and soil samples were air-dried, disaggregated, homogenized, and sieved to obtain the <1 mm fraction. All solid samples with the exception of HFOs were analyzed by multi-elemental inductively coupled plasma-mass spectrometry (ICP-MS) at Acme Analytical Laboratories, Vancouver (Canada). Accuracy and precision was monitored by repeated analyses of analytical and method blanks, duplicates, and standard reference materials. The detection limits were 0.5 and 0.1 mg/kg for As and Sb in the solid samples, respectively. Solid samples containing more than 2000 mg/kg Sb (upper detection limit at Acme Analytical Laboratories) were submitted for chemical analyses using AAS (Varian Spectr AA 220) equipped with a hydride generation system (VGA 76) at the accredited Geoanalytical Laboratories of the State Geological Institute of Dionyz Stur (Slovak Republic). Quality assurance, detection limits and other details of the chemical analyses can be found in Vrana et al. (1997). The accuracy of the sample analysis was checked by using NIST 2711 (Montana soil) reference material containing 105 ± 8 mg/kg As and 19.4 ± 1.8 mg/kg Sb. The measured values (100 ± 7 mg/kg for As and 17.3 ± 2.6 mg/kg for Sb; n = 3) were not significantly different from the certified values. The pHs of the solid samples were measured in water (solid:liquid ratio 1:2.5) using a WTW Multi 350i instrument. For the determination of the metalloid concentrations in HFOs, the samples were dissolved with 10 M HCl and the resulting extracts were analyzed at the accredited Geoanalytical Laboratories of the State Geological Institute of Dionyz Stur (Slovak Republic). To obtain samples for mineralogical characterization of the waste rocks and tailings, the collected material was pre-concentrated by panning in water and further separated in heavy liquids. Selected heavy concentrates were prepared for further study in the form of standard thin and polished sections, to be inspected in transmitted and reflected polarized light, respectively. Further information on the chemical composition of the mineral phases was obtained with a Cameca SX-100 electron microprobe (Bratislava) in wavelength-dispersive mode (WDS) under the conditions 20 kV, 20 nA, and beam size of 1–5 lm. The following lines, standards, and detectors were used for the determined elements: Mg (Ka, forsterite, TAP); Al (Ka, Al2O3, TAP); Si (Ka, SiO2, TAP); S, Fe, Cu (Ka, CuFeS2, PET); P (Ka, GaP, PET); Ca (Ka, wollastonite, PET); Mn (Ka, Mn, LIF); Co (Ka, Co, LIF); Ni (Ka, Ni, LIF); Zn (Ka, ZnS, LIF), Pb (Ma, PbS, PET); Sb (Lb, Sb2S3, PET), As (Lb, FeAsS,

E. Hiller et al. / Applied Geochemistry 27 (2012) 598–614

603

ˇ ucˇma mine site and locations of sampling points for waters, stream sediments, soils, hydrous ferric oxides and mine wastes. Fig. 6. Map showing the C

TAP). The counting time on each peak was 20 s, and on background 10 s (20 s for Sb). The mineralogy of HFOs was studied by X-ray diffraction (XRD) with a Bruker AXS D8 Advance diffractometer equipped with a Cu Ka radiation source and a diffracted-beam graphite monochromator. 3.3. Leaching tests Simple short-term column leaching tests were performed on selected samples of the mine wastes in the laboratory. Samples were maintained under oxic conditions to assess the potential risk of the release of As and Sb into the water. In addition to these studies, single ‘‘batch’’ experiments were also performed with stream sediments and soils. Columns with an internal diameter of 26 mm, length of 170 mm and an endpiece with a 0.45 lm filter were used.

The columns were uniformly packed with 60 g of the reactive materials (<1 mm fraction). Demineralized water with pH 6.6 was introduced into the column by a plastic tube connected to peristaltic pump which provided a constant flow rate. The downward flow was regulated at a rate of 5 mL/h for 5 days. During the experiment, five samples of leachate from each of the columns were collected. After determining the pH, temperature, and electrical conductivity, the leachates were stored at 4 °C for later analyses of As and Sb. Batch extraction experiments were performed using a 100 mL polyethylene centrifuge tubes. One gram of dry stream sediment or soil was shaken with 50 mL of demineralized water for 16 h, centrifuged at 4000 rpm for 20 min, and filtered through a 0.45 lm membrane filter. Concentrations of As and Sb in the column leachates and batch supernatants were determined by the same methods as described above for water samples. Standard ref-

604

E. Hiller et al. / Applied Geochemistry 27 (2012) 598–614

Table 1 Selected characteristics of the studied hydrothermal Sb deposits in the Western Carpathians (Slovakia).

Host rocks

Occurrence

Medzibrod

Poprocˇ

ˇ ucˇma C

Black shales and metabasic rocks (amphibolites and actinolitic schists) Quartz-carbonate lenses, veins, nests

Granites and granodiorites with amphibolite gneisses Quartz-carbonate veins, stockworks, shear zone 27000 t Sb 1943–1991 Stibnite

Migmatites, biotitic and two mica gneisses

Graphitic and sericitic phyllites, metapsammites, metarhyolitic tuffs Quartz-carbonate veins, impregnations 10300 t Sba Until 1965 Stibnite

Black phyllitic shales, porphyroids, metarhyolites Quartz-carbonate lenses, veins

Arsenopyrite Moderately warm, humid 900 mm/a Steep gorges, forest land 6.3–7.5

Arsenopyrite Moderately warm, humid 695–1135 mm/a Open forest, upward slopes 6.8–7.9

Unknown 1790–1922 Stibnite, berthierite

As minerals Climateb

Arsenopyrite Moderately warm, moderately humid 700 mm/a Steep valley, open forest 6.3–8.0

Water pH b

Dúbrava

Production Main exploitation period Sb minerals

Rainfallb Topography and vegetation

a

Pernek

Arsenopyrite Cool mountainous 850–1000 mm/a Steep, dense forest 6.4–8.2

Quartz-graphitic lenses, nests and veinlets 57000 t Sb ore 1938–1950 Stibnite, berthierite, gudmundite Arsenopyrite Moderately cool 900–1000 mm/a Relatively steep, open forest 7.0–8.1

272000 t Sb ore 1918–1952 Stibnite

Between 1931 and 1965. According to the Landscape Atlas of the Slovak Republic (2002).

erence material NIST 1643e (simulated freshwater) was used to check the quality of the As and Sb determinations in the leachates and were within 6% for both elements. 4. Results 4.1. Pernek The water sample collected above the waste-rock dumps was of the Ca–Mg–HCO3–SO4 type and its pH was 7.7 (Table 2). The composition of this water changed to SO4 type below the waste-rock dumps due to interaction with waste rocks containing mainly pyrite. Only mine water discharging from the Pavol adit was slightly acidic (pH = 6.3) with a SO4-dominated composition. Waters collected downstream of the mines remained of the SO4 type, but they had neutral pH values. Groundwater extracted from household wells contained mainly HCO3 ions followed by Ca. Stream waters had low and constant As concentrations (Fig. 7). The lowest Sb concentration was recorded either in stream water upstream of the mines or in groundwater, while the highest Sb concentration (31 lg/L) was in water draining from waste rock dumps. The Sb content of waters decreased downstream likely due to adsorption of Sb on abundant HFOs. HFO precipitates contained up to 2009 mg/kg As and 2155 mg/kg Sb (Table 3). The concentrations of As and Sb in the stream sediments are given in Table 3. The former mining activities at Pernek have led to high As and Sb concentrations in stream sediments. The sediments of the right-hand tributary of the Kostolny´ Creek contained up to 390 mg/kg As and 703 mg/kg Sb. Lower As and Sb concentrations were measured in the stream sediment of the Kostolny´ Creek (PeS-5, Fig. 2) since no mining operations were undertaken along the creek. However, the stream sediment concentrations greatly exceeded the maximum acceptable value of 15 mg/kg for Sb (Anon, 1998) and also for As (55 mg/kg) with the exception of two samples. Forest soils sampled at the site had typical acidic pH values (4.5–6.2) and exhibited high concentrations of both elements (Table 3).

and neutral, although molar Ca/Mg and HCO3/SO4 ratios decreased compared to the upstream water. This shift in stream water composition is due to mine waters which drain into the Paludzˇanka Creek and have much lower molar Ca/Mg and HCO3/SO4 ratios, even though they are neutral to slightly alkaline. Groundwaters collected in alluvial sediments below the impoundments were slightly acidic to neutral and two out of 4 samples contained more SO4 than HCO3. Waters discharging from adit entrances had very high Sb concentrations, up to 9300 lg/L, and also elevated As values (Table 2 and Fig. 7). These high Sb concentrations in discharge waters are not attenuated by adsorption onto HFOs as no HFOs were observed in the vicinity of adit entrances. Only waters emanating from the oldest impoundment are precipitating HFOs (DuO-1–7, Fig. 3), which efficiently adsorbed both metalloids as reflected by the high As and Sb concentrations (Table 3). Stream water in the Paludzˇanka Creek above the mines contained 10 lg/L Sb and 9 lg/L As. Downstream of the confluence of mine waters with the Paludzˇanka Creek, stream waters were enriched in Sb above the drinking water limit, even below the village of Lazisko, which is located almost 6 km NE of this junction (Figs. 3 and 7). Arsenic in stream water was more attenuated and its content stayed close to or below the drinking water limit. Most of the groundwater samples taken from alluvial sediments and household wells also contained elevated Sb (4–126 lg/L), but not As. Upstream of the mines, stream sediment in the Paludzˇanka Creek had elevated As and Sb concentrations (Table 3). This is evidence that sulfide mineralization also extends beyond the abandoned mines. Downstream of the mines, As and Sb concentrations of stream sediments increased significantly up to 140 mg/kg and 644 mg/kg, respectively. The Slovak average soil As and Sb concentrations are 10.4 and 1.8 mg/kg, respectively (Cˇurlík and Šefcˇík, 1999). Most of the soils had As and Sb concentrations that exceeded these respective average soil values many times as well as the estimated global average soil contents (Reimann and de Caritat, 1998). The highest concentrations of As and Sb were in soils located close to Paludzˇanka Creek (DuP-3 and -10, Fig. 3). The soils were mostly acidic (Table 3), and their pH values correlated positively with total Ca content (r2 = 0.76, p < 0.01).

4.2. Dúbrava 4.3. Medzibrod Stream water above the mine site was neutral and Ca2+ and HCO3 ions were dominant dissolved components. Major ion chemistry of this water downstream of the mines continued to be Ca–HCO3 type

All water samples at Medzibrod were of the Ca–Mg–HCO3–SO4 type and had pH values ranging from 7.0 to 8.1. Sulfate concentra-

605

E. Hiller et al. / Applied Geochemistry 27 (2012) 598–614 Table 2 Description of water character and selected chemical analyses of water samples. Sample

a b

Character

pH

Eh

Ca

mV

mg/L

Mg

HCO3

SO4

Type

Fe

As

Sb

lg/L

Pernek PeH-1 PeH-2 PeH-3 PeH-4 PeH-5 PeH-6 PeH-7 PeH-8 HW-1a HW-2

Upstream water Downstream water Adit water Downstream water Downstream water Downstream water Downstream water Downstream water Groundwater Groundwater

7.7 6.9 6.3 7.1 7.5 7.7 7.7 8.0 7.3 7.1

449 469 297 279 436 450 463 435 368 385

29 72.6 120 103 97.6 76.7 66.5 63.6 135 141

5.8 25.1 42.4 36.0 33.3 23.6 18.7 10.1 9.66 19.4

73.8 49.2 64.6 52.2 49.2 67.6 89.1 160 297 382

62.7 259 488 434 377 261 204 89.5 92.9 71.9

Ca–Mg–SO4–HCO3 Ca–Mg–SO4 Ca–Mg–SO4 Ca–Mg–SO4 Ca–Mg–SO4 Ca–Mg–SO4 Ca–Mg–SO4–HCO3 Ca–HCO3–SO4 Ca–HCO3–SO4 Ca–HCO3

94.0 247 6940 673 105 95.0 124 43.0 56.0 21.0

1.0 1.0 5.0 1.0 <1.0 1.0 1.0 2.0 1.0 3.0

1.0 31.0 14.0 15.0 15.0 11.0 9.0 3.0 <1.0 1.0

Dúbrava DuH-1 DuH-2 DuH-3 DuH-4 DuH-5 DuH-6 DuH-7 DuH-8 DuH-9 DuH-10 DuH-11 DuH-12 BH-1b BH-2 BH-3 BH-4 BH-5 HW-1

Upstream water Adit water Adit water Adit water Adit water Adit water Adit water Downstream water Impoundment water Spring water Downstream water Downstream water Groundwater Groundwater Groundwater Groundwater Groundwater Groundwater

7.3 8.2 8.0 7.6 8.1 8.2 8.2 7.6 8.0 8.0 7.4 7.4 6.4 6.4 6.4 7.0 7.8 7.4

530 464 484 457 463 504 483 466 378 448 496 474 375 468 415 348 569 –

5.4 52.2 42.7 23.3 47.7 35.6 44.1 7.3 54.9 48.2 8.6 10.7 63.6 19.2 66.5 33.8 32.1 28.3

1.2 33.1 21.9 12.2 25.2 17.9 20.2 2.4 18.6 28.6 2.6 3.8 26.1 6.6 24.2 23.9 17.1 10.9

12.3 141 111 67.7 105 141 126 18.4 200 267 18.4 33.8 36.2 43.0 138 132 129 104

5.30 142 100 48.8 137 93.1 119 11.1 48.4 12.8 10.5 12.5 196 33.6 144 65.4 85.8 22.2

Ca–Mg–HCO3–SO4 Mg–Ca–SO4–HCO3 Ca–Mg–SO4–HCO3 Ca–Mg–HCO3–SO4 Ca–Mg–SO4–HCO3 Ca–Mg–HCO3–SO4 Ca–Mg–SO4–HCO3 Ca–Mg–HCO3–SO4 Ca–Mg–HCO3–SO4 Ca–Mg–HCO3 Ca–Mg–HCO3–SO4 Ca–Mg–HCO3–SO4 Ca–Mg–SO4 Ca–Mg–HCO3–SO4 Ca–Mg–SO4–HCO3 Mg–Ca–HCO3–SO4 Ca–Mg–HCO3–SO4 Ca–Mg–HCO3–SO4

40.0 46.0 51.0 504 39.0 145 41.0 34.0 241 38.0 36.0 34.0 695 303 306 184 211 35.0

9.0 62.0 34.0 57.0 39.0 34.0 16.0 10.0 16.0 2.0 10.0 7.0 3.0 3.0 3.0 4.0 7.0 <1.0

10.0 9300 3280 1650 1300 830 1330 79.0 104 4.0 76.0 76.0 107 30.0 4.0 45.0 20.0 126

Medzibrod MdH-1 MdH-2 MdH-3 MdH-4 MdH-5 MdH-6 BH-1

Upstream water Dump water Adit water Dump water Impoundment water Downstream water Groundwater

7.1 7.8 7.0 7.4 8.1 7.9 7.6

496 455 353 435 450 419 413

13.1 77.7 58.0 63.1 21.1 26.8 60.3

5.2 30.5 32.8 25.7 4.86 9.32 23.2

46.1 279 232 246 97.6 107 279

15.2 96.0 108 77.8 30.9 24.8 18.2

Ca–Mg–HCO3–SO4 Ca–Mg–HCO3–SO4 Ca–Mg–HCO3–SO4 Ca–Mg–HCO3–SO4 Ca–Mg–HCO3–SO4 Ca–Mg–HCO3–SO4 Ca–Mg–HCO3

106 175 112 208 69.0 82.0 121

32.0 255 180 204 100 90.0 7.0

11.0 1290 445 870 110 128 32.0

Poprocˇ PoH-1 PoH-2 PoH-3 PoH-4 PoH-5 PoH-6 PoH-7 PoH-8 PoH-9 HW-1 HW-2 PoH-10 PoH-11 BH-1 BH-2

Upstream water Adit water Downstream water Adit water Mine water Downstream water Downstream water Impoundment water Downstream water Groundwater Groundwater Spring water Downstream water Groundwater Groundwater

6.9 6.9 7.0 6.3 6.6 6.8 7.7 6.2 7.4 7.0 7.6 6.1 7.5 6.5 6.4

386 354 338 214 370 197 400 256 272 305 433 472 261 338 425

8.8 24.5 12.8 45.1 20.3 23.3 32.1 49.2 25.3 46.1 77.2 13.3 29.9 77.8 40.8

3.7 14.5 6.1 36.5 9.6 14.8 17.2 51.5 12.8 12.0 14.0 2.9 14.8 34.1 8.3

24.6 111 39.9 135 56.1 55.0 73.2 95.2 64.4 171 295 21.4 82.9 184 67.5

23.2 49.8 28.4 240 45.5 86.1 87.7 313 76.0 31.9 25.3 38.2 67.4 164 41.9

Ca–Mg–SO4–HCO3 Ca–Mg–HCO3–SO4 Ca–Mg–HCO3–SO4 Mg–Ca–SO4–HCO3 Ca–Mg–SO4–HCO3 Mg–Ca–SO4–HCO3 Ca–Mg–SO4–HCO3 Mg–Ca–SO4 Ca–Mg–SO4–HCO3 Ca–Mg–HCO3 Ca–HCO3 Ca–SO4–HCO3 Ca–Mg–SO4–HCO3 Ca–Mg–SO4–HCO3 Ca–Mg–HCO3–SO4

153 245 151 32,700 271 4630 401 8190 324 74.0 288 67.0 853 250 229

5.0 5.0 7.0 2150 9.0 197 28.0 1110 50.0 1.0 4.0 2.0 27.0 78.0 3.0

36.0 750 230 600 700 335 390 175 500 8.0 35.0 8.0 440 1000 5.0

Cˇucˇma CuH-1 CuH-2 CuH-3 CuH-4 CuH-5 CuH-6 CuH-7 CuH-8 BH-1 BH-2 HW-1 HW-2 CuH-9 CuH-10 CuH-11

Adit water Downstream water Upstream water Adit water Downstream water Impoundment water Downstream water Downstream water Groundwater Groundwater Groundwater Groundwater Spring water Water resource Water resource

7.7 7.3 7.4 7.1 7.9 7.9 7.7 7.6 6.8 7.6 6.5 6.3 6.3 7.3 7.3

398 512 498 238 451 490 457 478 285 501 380 527 487 491 222

44.7 15.2 6.7 98.8 18.5 62.9 21.1 19.8 67.2 60.1 20.6 16.7 8.31 9.12 39.4

19.7 4.3 1.5 62.8 7.1 66.8 9.8 8.2 27.7 59.1 5.6 5.0 4.1 2.8 15.4

166 43.0 21.2 611 61.4 295 82.9 70.6 184 246 48.0 49.0 18.1 24.2 142

70.8 29.5 2.70 67.5 23.8 217 36.4 33.3 174 193 15.4 23.2 8.60 6.60 47.9

Ca–Mg–HCO3–SO4 Ca–Mg–HCO3–SO4 Ca–Mg–HCO3 Mg–Ca–HCO3 Ca–Mg–HCO3–SO4 Mg–Ca–HCO3–SO4 Ca–Mg–HCO3–SO4 Ca–Mg–HCO3–SO4 Ca–Mg–SO4–HCO3 Mg–Ca–HCO3–SO4 Ca–Mg–HCO3–NO3 Ca–Mg–HCO3–SO4 Ca–Mg–NO3–HCO3 Ca–Mg–HCO3–NO3 Ca–Mg–HCO3–SO4

183 790 39.0 555 40.0 62.0 419 613 5850 88.0 117 93.0 46.0 57.0 83.0

84 7.0 6.0 1350 670 110 50 15 285 6.0 5.0 1.0 1.0 1.0 27

3540 180 2.0 17 116 1060 310 260 970 194 13 86 1.0 1.0 170

HW – water from household wells. BH – water extracted from boreholes.

606

E. Hiller et al. / Applied Geochemistry 27 (2012) 598–614

Fig. 7. Schematic examples of variations in As and Sb concentrations of waters downstream at the Pernek, Dúbrava, Poprocˇ and Cˇucˇma mine sites.

tion in stream water above the mines was lower than in this water downstream after mixing with mine waters. Uncontaminated water of the Borovsky´ Creek upstream of the mines had elevated concentrations of both metalloids (Table 2). The As and Sb contents in stream water below the mines increased by 3- and 10-times, respectively, as a result of the mixing with mine and spring waters having strongly elevated As (up to 255 lg/L) and Sb (1290 lg/L) concentrations. HFOs collected in

water discharging from the Murgaš adit (Fig. 4) had As concentrations up to 205,882 mg/kg. This indicates that HFO deposits play an important role in attenuation of As in discharge waters. Groundwater from a borehole in the stone-pit contained elevated Sb exceeding the drinking water limit of 5 lg/L (WHO, 2006). In Borovsky´ Creek stream sediment close to the Murgaš adit contained 9180 mg/kg As and 890 mg/kg Sb. Both, As and Sb concentrations in sediments decreased with distance from the mine

607

E. Hiller et al. / Applied Geochemistry 27 (2012) 598–614 Table 3 Arsenic and Sb concentrations in stream sediments, soils and hydrous ferric oxides (HFOs) collected at the studied mine sites. Sediment

As

Sb

Soil

mg/kg

a b

As

Sb

pH

mg/kg

Ca

HFO

(wt.%)

As

Sb

mg/kg

Pernek PeS-1 PeS-2 PeS-3 PeS-4 PeS-5 PeS-6 PeS-7

65 351 217 244 45 390 49

56 703 439 583 48 96 70

PeP-1 PeP-2 PeP-3 PeP-4 PeP-5 PeP-6

394 364 514 421 324 100

187 227 894 632 479 121

4.5 5.2 5.1 6.0 m.ma 6.2

2.05 1.91 1.86 2.04 2.24 2.52

PeO-1 PeO-2 PeO-3 PeO-4 PeO-5 PeO-6 PeO-7 PeO-8 PeO-9 PeO-10

Dúbrava DuS-1 DuS-2 DuS-3 DuS-4 DuS-5

25 45 140 84 99

46 72 326 547 644

DuP-1 DuP-2 DuP-3 DuP-4A DuP-4B DuP-5 DuP-6 DuP-7A DuP-7B DuP-8 DuP-9 DuP-10A DuP-10B DuP-11

25 28 187 66 74 40 8.9 9.8 8.3 17 9.1 845 930 146

29 224 1249 89 47 68 19 9.0 4.8 9.8 11 8739 9619 525

4.1 5.9 6.3 4.4 4.8 4.1 3.8 4.8 5.1 5.0 5.8 5.4 5.7 6.5

0.21 1.17 0.90 0.18 0.19 0.30 0.21 0.23 0.21 0.27 0.69 0.58 0.19 0.89

Medzibrod MdS-1 MdS-2 MdS-3 MdS-4 MdS-5

9180 30 648 513 199

890 32 299 268 93

MdP-1 MdP-2 MdP-3 MdP-4

13 10,250 562 44

2.0 723 793 21

7.6 n.m n.m 6.6

Poprocˇ PoS-1 PoS-2 PoS-3 PoS-4 PoS-5

52 5560 634 292 171

373 1244 1360 245 215

PoP-1 PoP-2 PoP-3A PoP-3B PoP-4 PoP-5 PoP-6 PoP-7 PoP-8 PoP-9

636 37 1714 1764 316 108 28 57 155 65

6786 13 3079 3291 694 143 33 40 1989 51

Cˇucˇma CuS-1 CuS-2 CuS-3 CuS-4 CuS-5 CuS-6 CuS-7

52 187 75 73 94 149 80

11 320 20 552 581 600 131

CuP-1 CuP-2 CuP-3 CuP-4 CuP-5 CuP-6 CuP-7 CuP-8

30 133 41 73 83 70 190 159

6.2 201 10 466 401 194 512 782

1024 2009 1005 695 820 697 687 918 573 606

1187 1602 2109 1237 1047 2072 1647 2155 1075 1371

DuO-1 DuO-2 DuO-3 DuO-4 DuO-5 DuO-6 DuO-7

10,864 12,140 13,257 6500 9825 5925 7700

1440 847 3615 3963 1375 1025 1425

13.6 n.m n.m 0.29

MdO-1 MdO-2

202,500 205,882

7750 10,294

4.5 4.4 3.0 3.2 6.6 4.6 4.2 5.3 4.9 5.0

0.16 0.15 0.02 0.02 1.05 0.19 0.14 0.09 0.14 0.09

PoO-1 PoO-1b PoO-2 PoO-2b PoO-2b

28,401 8600 52,809 34,000 60,412

17,173 5200 12,360 6100 10,838

5.3 5.5 4.2 5.3 4.8 4.2 7.5 5.1

1.59 0.64 0.41 0.95 0.86 0.41 1.12 0.26

CuO

23,860

8926

Not measured. Samples taken at the same site.

workings, although the sediment farthest from the mines (MdS-5, Fig. 4) had high contents of both metalloids, greatly exceeding their respective maximum acceptable values. Soils were circum-neutral due to a high carbonate content and had variable As and Sb concentrations (Table 3). Soil sampled close to the Murgaš adit (MdP-1) displayed the highest As content followed by soil below the impoundment, which had the highest Sb content (MdP-3). 4.4. Poprocˇ The sample from the Olšava Creek upstream of the mine workings was characterized by an elevated SO24 concentration, but had circum–neutral pH. However, surface water of the Olšava Creek

downstream was clearly affected by mine waters as indicated by an increase in SO24 and Mg2+ contents, although pH values remained circum–neutral. Waters discharging from adits and impoundments were slightly acidic (pH = 6.1) and dominated by Mg2+ and SO24 ions. On the other hand, groundwater collected from two household wells at the upper end of the village of Poprocˇ (Table 2, Fig. 5) was of Ca–HCO3 type and neutral. Waters draining from adits and impoundments had high concentrations of both metalloids (up to 2150 lg/L As and 750 lg/L Sb). This was also true for Sb concentrations in waters of Olšava Creek, which showed no decrease in the studied section of the creek (Fig. 7). These mine waters are characterized by the precipitation of abundant HFO layers. HFOs contained 1.7- to 5.6-times

608

E. Hiller et al. / Applied Geochemistry 27 (2012) 598–614

more As than Sb (Table 3). HFOs at this mine site likely play a significant role in the natural attenuation of both metalloids in stream waters, although their significance was more pronounced for As than Sb. For example, discharge water from the Agnes adit had 2150 lg/L As, which rapidly decreased by a factor of 10, whereas Sb content was decreased by only a half. Groundwater from household wells contained Sb (8–35 lg/L) well above the drinking water limit and groundwater close to the impoundment (BH-1) displayed a much higher content of Sb than As (Fig. 7). Sediments in the Olšava Creek below the adits and mine workings had As and Sb levels up to 5560 mg/kg and 1360 mg/kg, respectively. Although there was a decrease in sediment As and Sb concentrations with distance from the main mine workings, stream sediment near the village of Jasov (situated 5 km south of the mines, Fig. 5) still had high As and Sb concentrations (PoS-

5). Peak As and Sb concentrations were observed in the soils close to either the adits (PoP-1) or the impoundments (PoP-3, -4, and -8). However, soils collected from steep slopes of the valley several tens of meters above the mine workings had much lower contents of both metalloids (PoP-2, -6, -7, and -9), but were still elevated when compared to the Slovak average soil values (Table 3). All of the soils were characterized by acidic pH values, which correlated positively to total soil Ca content (r2 = 0.77, p < 0.01). 4.5. Cˇucˇma Water in Laz Creek upstream of the first open adit (CuH-3) was of Ca–Mg–HCO3 type with a low SO24 content and circum–neutral pH. Major ion chemistry of the creek waters downstream of the mines and their pH values remained relatively unchanged,

Fig. 8. Back-scattered electron images of the sulfide minerals and their weathering products: (a and c) arsenopyrite (FeAsS) and their weathering rims in samples DuT and MdT-A, respectively, (b and d) pyrite (FeS2) and the weathering rims in sample DuT and PeD-3, respectively, (e) Fe oxide rich in Sb (white) cementing quartz crystals (dark gray) in sample PoT-3, (f) Fe oxide in sample PoT-1, (g) pseudomorphic replacement of pyrite grain by Fe oxides in sample PeD-3, (h) Fe oxides (light gray coatings) enclosing carbonate grain (dark gray) in sample CuT-A.

609

E. Hiller et al. / Applied Geochemistry 27 (2012) 598–614

although some mine waters draining into the main streams were characterized by the dominance of Mg2+ ions. Three water samples ˇ ucˇma had elevated NO3 contents and slightly in the village of C acidic pH values, thus causing a shift from Ca–Mg–HCO3–SO4 to a Ca–Mg–HCO3–NO3 dominated composition. This change in major ion water chemistry is due to improper storage of barnyard manure and dung-water by people living in the village. Upstream of the mines, Laz Creek water contained low concentrations of Sb and As (Fig. 7, Table 2). These metalloid concentrations in stream water below the adits increased rapidly to 116 lg/L Sb and 670 lg/L As. Farther downstream of the confluence of the discharge impoundment water with the Laz Creek, the Sb concentration in water continuously increased to 310 lg/ L, whereas the As concentration decreased rapidly to 50 lg/L (Fig. 7). The contamination status of this water was slightly chanˇ ucˇmiansky Creek (CuH-8, ged below the confluence with the C Fig. 6). The As and Sb concentrations in water decreased, but the decrease was more significant for As than Sb. This change was ˇ ucˇmiansky Creek, which likely due to dilution with water of the C

displayed low As content, but elevated Sb content (CuH-2). All of the mine waters had high Sb and As concentrations of up to 3540 lg/L and 1350 lg/L, respectively (Fig. 7). The Sb concentrations found in groundwater from household wells might be of concern for public health as they were far above the drinking water limit (Fig. 7). The former mining activities were situated mainly along the Laz ˇ ucˇmiansky Creeks. Although there is some indication of eleand C vated As and Sb concentrations in sediments either above or away from the main mine workings (CuS-1 and -3, Table 3), contamination of the sediments with As and Sb occurred mostly below the adits and impoundment. Maximum As and Sb concentrations found in stream sediments were 187 mg/kg and 600 mg/kg, respectively. Soils at the mine site had greatly elevated contents of As and Sb with a maximum of 190 mg/kg for As and 782 mg/ kg for Sb (Table 3). Although there was a clear input from the mine workings, it was also apparent that elevated soil contents of As and Sb occurred above the mine workings (CuP-1) and away from the mines (CuP-3). This suggests the presence of mineralization up-

Table 4 Representative electron microprobe analyses of the weathering products of the primary sulfide minerals (wt.%) in selected samples of the waste-rock dump (Pernek) and the flotation tailings (Dúbrava, Medzibrod, Poprocˇ and Cˇucˇma) materials. Sample Rims on pyrite PeD-3

Fe

As

Sb

Ca

Mn

Pb

S

Si

Al

Total

55.02 47.07 53.97 49.19 40.49 42.01 56.05 55.86 58.01

0.00 3.66 0.04 0.62 2.54 0.20 0.00 0.00 0.03

0.08 0.52 0.39 1.00 11.94 0.09 0.05 0.21 0.15

0.38 1.26 1.84 0.54 0.64 0.11 0.05 0.34 0.28

0.00 4.02 0.00 0.37 0.12 1.15 2.74 1.89 0.63

0.04 0.05 0.04 0.14 0.07 0.04 0.07 0.05 0.00

0.45 0.35 0.55 0.45 0.02 0.52 0.10 0.67 3.05

4.55 1.82 3.48 2.30 1.95 0.64 0.11 0.10 0.06

0.02 0.84 0.00 0.22 1.23 0.19 0.00 0.03 0.00

90.47 88.84 81.83 74.73 85.89 66.27 85.65 88.65 96.09

Rims on arsenopyrite PeD-3 22.36 24.71 DuT 28.37 46.48 42.09 CuT-A 50.24 48.80 49.17

24.45 31.53 19.87 2.89 14.40 12.66 10.31 10.38

1.17 0.41 1.54 6.97 1.55 0.16 0.19 0.13

1.10 0.95 4.94 1.23 2.10 1.86 1.62 1.54

0.02 0.01 0.15 0.20 0.03 0.00 0.00 0.00

0.32 0.26 0.09 0.14 0.05 0.03 0.06 0.01

0.18 0.07 0.22 0.26 4.24 0.14 0.11 0.24

2.43 0.45 1.11 1.83 0.63 0.17 0.16 0.16

2.93 0.88 0.14 0.41 0.06 0.01 0.02 0.01

84.74 89.11 75.68 80.19 90.61 96.72 90.77 91.36

Fe oxides rich in Sb PeD-3 18.55 22.00 DuT 23.88 22.69 MdT-A 12.26 8.22 16.03 MdT-B 34.35 33.51 PoT-2 23.44 31.09 PoT-4 15.03 16.61 23.45

0.46 0.24 0.96 0.33 2.66 4.88 3.63 3.90 3.89 3.22 2.14 2.46 2.45 5.26

36.28 33.77 40.91 43.04 23.97 37.91 35.09 20.95 20.45 22.81 28.23 29.08 32.56 22.11

0.92 1.22 0.45 0.31 2.50 6.08 3.47 1.39 1.38 0.11 0.22 0.40 0.36 0.43

0.01 0.00 0.02 0.02 0.03 0.02 0.03 0.23 0.09 0.00 0.00 0.07 0.09 0.00

0.12 0.14 5.13 4.79 0.96 0.25 0.19 0.76 0.69 1.88 3.65 3.16 3.08 1.10

0.28 0.40 0.05 0.09 0.01 0.01 0.01 0.04 0.07 0.03 1.18 0.23 0.63 1.53

0.65 0.74 0.34 0.43 7.40 0.72 0.46 0.09 0.10 0.74 0.40 1.25 0.93 2.13

0.05 0.08 0.01 0.07 4.97 0.17 0.20 0.14 0.08 0.14 0.16 0.20 0.04 0.26

79.22 81.61 88.77 88.78 84.17 80.99 82.29 87.42 85.28 84.13 83.04 75.52 79.15 82.23

0.61 0.22 0.34 0.28 21.81 0.66 0.04 0.07 9.88 1.99 0.11 0.00

0.11 0.06 3.63 5.07 1.76 0.36 0.01 0.00 11.52 7.48 0.15 0.01

0.07 0.05 0.33 0.47 1.19 0.22 0.09 0.04 0.40 0.10 0.02 0.02

0.00 0.00 0.42 0.41 0.00 0.10 0.20 0.24 0.00 0.09 0.40 0.39

0.03 0.07 0.05 0.11 0.14 0.05 0.01 0.03 0.35 0.53 0.06 0.02

0.21 0.20 0.02 0.05 0.44 0.24 0.15 0.05 1.23 0.46 0.01 0.01

1.15 1.53 1.92 1.74 1.50 0.04 0.25 0.67 2.06 1.51 1.32 1.33

0.09 0.00 0.01 0.06 0.71 0.00 0.02 0.00 0.73 0.22 0.14 0.12

92.26 92.36 81.20 80.78 78.99 84.75 86.22 86.81 92.01 85.91 85.94 85.56

DuT MdT-A CuT-B CuT-A

Fe oxides PeD-2 DuT MdT-A PoT-1 PoT-3 CuT-B

61.00 61.05 54.50 52.74 24.67 56.95 58.94 59.09 35.85 46.81 56.56 56.69

610

E. Hiller et al. / Applied Geochemistry 27 (2012) 598–614

stream and beyond the immediate mine workings. Soil pH values were variable, but mostly acidic. 4.6. Mineralogical characterization of mine wastes The sampled waste rocks and tailings contained the primary mineral assemblage of the mineralization as well as the host rocks. The most abundant ore minerals identified in heavy fractions of the mine wastes were pyrite and arsenopyrite. Antimony sulfides such as stibnite and berthierite were rare. The ore minerals present in mine wastes showed signs of oxidative weathering. Pyrite, arsenopyrite, and berthierite were replaced by weathering rims (Fig. 8a–d) or

totally altered (Fig. 8e–h). Both rims and alteration products were of variable composition generally enriched in Fe, As, Sb, and rarely Pb (Table 4). The rims on arsenopyrite had high As contents (up to 31.5 wt.%), but also increased amounts of Sb (up to 7.0 wt.%) exceeding its primary content in arsenopyrite. Similarly, the rims developed on pyrite contained up to 29.9 wt.% As and 11.9 wt.% Sb in some cases, although more frequently exhibited low As and Sb contents, as expected for a mineral depleted in both metalloids. This was also the case of the rims on berthierite from the Poprocˇ tailings, a mineral with As contents ranging between 0.19 and 0.56 wt.% (Klimko et al., 2009), containing elevated As (up to 3.3 wt.%) and up to 43.6 wt.% Sb. The observed enrichment of the weathering rims in

Table 5 Concentrations of As and Sb in mine wastes, stream sediments and soils, and their concentrations in the leachates. As

Sb

Sbtot/Astota

a

Sb

Sb/Asb

lg/L

mg/kg

b

As

As

Sb

% of total

Mine wastes PeD-1 PeD-2 PeD-3 DuT MdT-A MdT-B PoT-1 PoT-2 PoT-3 PoT-4 CuT-A CuT-B

574 63.4 933 178 5111 5166 1497 1128 2342 1289 27.0 45.0

365 16.9 1596 762 2807 1802 3190 1990 1509 3847 9861 161

0.64 0.27 1.68 4.28 0.55 0.35 2.13 1.76 0.64 2.98 365 3.58

5–7 27–46 22–93 11–15 3130–7013 2690–8400 4–17 32–40 72–316 30–39 2–10 3–6

7–39 7–104 4–417 428–1039 1876–4060 2048–4050 20–34 255–353 257–716 77–137 1976–4039 2–6

2.70 1.56 2.12 54 0.59 0.58 2.36 8.14 3.08 2.63 580 0.88

0.01 0.53 0.05 0.08 0.68 0.57 0.007 0.03 0.06 0.03 0.19 0.09

0.05 3.10 0.06 0.94 0.73 0.95 0.008 0.15 0.29 0.02 0.29 0.02

Stream sediments PeS-1 PeS-2 PeS-3 PeS-4 DuS-1 DuS-2 DuS-3 DuS-4 DuS-5 MdS-1 MdS-2 MdS-3 MdS-4 MdS-5 PoS-1 PoS-2 PoS-3 PoS-4 PoS-5 CuS-1 CuS-3 CuS-5 CuS-6 CuS-7

65 351 217 244 25 45 140 84 99 9180 30 648 513 199 52 5560 634 292 171 52 75 94 149 80

56 703 439 583 46 72 326 547 644 890 32 299 268 93 373 1244 1360 245 215 11 20 581 600 131

0.87 2.00 2.02 2.39 1.82 1.59 2.33 6.51 6.51 0.10 1.06 0.46 0.52 0.47 7.18 0.22 2.15 0.84 1.26 0.21 0.26 6.18 4.03 1.64

20 56 37 50 22 39 58 35 31 1960 112 390 319 126 9.0 89 71 39 12 42 27 37 47 25

76 214 133 138 42 104 345 317 597 1010 42 121 88 44 427 331 829 273 307 5.0 12 301 300 150

3.80 3.82 3.59 2.76 1.92 2.66 5.99 8.96 19.0 0.52 0.38 0.31 0.28 0.35 47.4 3.72 11.7 6.94 24.9 0.12 0.44 8.14 6.38 6.00

1.54 0.80 0.85 1.02 4.40 4.33 2.06 2.10 1.58 1.07 18.6 3.01 3.11 3.16 0.87 0.08 0.56 0.67 0.36 4.04 1.80 1.97 1.58 1.56

6.75 1.52 1.51 1.18 4.64 7.24 5.29 2.89 4.63 5.67 6.66 2.03 1.64 2.34 5.72 1.33 3.05 5.57 7.15 2.29 3.06 2.59 2.50 5.72

Soils PeP-1 PeP-2 PeP-3 PeP-4 PeP-5 PeP-6 PoP-1 PoP-2 PoP-3A PoP-4 PoP-5 PoP-8 PoP-9

394 364 514 421 324 100 636 37 1714 316 108 155 65

187 227 894 632 479 121 6786 13 3079 694 143 1989 51

0.47 0.62 1.74 1.50 1.48 1.21 10.7 0.37 1.80 2.20 1.33 12.9 0.78

30 31 56 26 13 12 133 9.0 71 79 42 34 12

30 35 110 145 26 25 871 7.0 93 343 89 500 78

1.00 1.13 1.98 5.69 2.08 2.17 6.55 0.78 1.32 4.34 2.14 14.9 6.46

0.15 0.17 0.22 0.12 0.08 0.23 0.42 0.49 0.08 0.50 0.77 0.43 0.37

0.32 0.31 0.25 0.46 0.11 0.41 0.26 1.04 0.06 0.99 1.24 0.50 3.06

Mass ratio of Sb to As in the solids. Mass ratio of Sb to As in the leachates.

E. Hiller et al. / Applied Geochemistry 27 (2012) 598–614

elements originally not present in the parent ore minerals suggested that the pore waters were able to carry the elements in their dissolved forms to the forming rims on sulfide minerals. Iron oxides with essentially no As and Sb were probably derived from the pseudomorphic replacement of pyrite grains since inclusions of pyrite were still observed in these mineral phases. On the other hand, a large part of the secondary oxides present was probably formed by precipitation reactions from the pore waters present in the tailings and waste-rock materials. These As- and Sb-bearing oxides were found with the following textures: individual grains with alotriomorphic and hypidiomorphic boundaries, grains with obvious zonality or no zonality, fillings associated with carbonates and silicates, framboidal aggregates, and aggregates cementing quartz crystals (Fig. 8). The electron microprobe analyses of secondary oxides showed that Fe oxides often contained significant amounts of As (up to 21.8 wt.%) (Table 4). Antimony was mostly present in the form of Sb, Sb(Fe) and Fe(Sb) oxides. These oxides might derive from complete replacement of Sb ores or precipitation from solutions which circulate through the impoundments. The uppermost layers of the tailings had a yellowish-red color and below this zone, the tailings consisted of gray sediments. The mine wastes were characterized by variable concentrations of metalloids that ranged from 27 to 5166 mg/kg As and from 16.9 to 9861 mg/kg Sb (Table 5).

611

with upstream waters and various small streams draining into the creeks whose net flow rate exceeds that of mine waters. 5.2. Metalloids in waters Within the studied mines, the lowest As and Sb concentrations were measured in waters of the Pernek (Table 2, Fig. 7). This was related to the minerals present in waste rocks and mineralized rocks remaining after exploitation of Sb-ores. At Pernek, only pyrite was identified as a common sulfide mineral in the rocks, followed by minor arsenopyrite. Although there were Fe oxides containing up to 33.8 wt.% Sb in waste rocks (Table 4), these minerals comprised only an insignificant proportion of the bulk samples. In contrast, stream waters of the other 4 mine sites had much higher As and Sb concentrations than those of Pernek. One of the possible explanations is that mineralized rocks and tailings at these sites still contain a relatively large amounts of As- and Sb-bearing sulfides and other mineral phases enriched in both metalloids. Mineralogical studies revealed that all sampled tailings at the 4 mine sites were characterized by the abundance of arsenopyrite, and Fe-oxides rich in Sb (Table 4) were also much more abundant than at Pernek. Moreover, unlike the Pernek site, stibnite is a common sulfide in the reduced zone of the tailings impoundments left at the 4 mine sites (Maruška et al., 2000; Lalinská and Chovan, 2006; Klimko et al., 2009).

5. Discussion 5.1. Major ion chemistry Calcium and HCO3 dominated the composition of the waters upstream of mine workings and those extracted from household wells. All of these waters also displayed neutral pH values. This is characteristic of the dissolution of Ca-bearing carbonates, which are present in veins in the studied areas. The molar Ca/Mg ratios of the upstream waters were close to 3:1, except those from the Medzibrod and Poprocˇ with molar Ca/Mg ratios of about 3:2. The difference in Ca/Mg ratios between the mine sites indicates that major ion chemistry of the upstream waters is dependent on dissolution of a variety of Ca- and Mg-bearing carbonates, which are present in different proportions at the studied mine sites. Distinct major ion composition from that of the upstream waters was found in most of the mine waters (Table 2). In the Pernek, Dúbrava, and Poprocˇ areas, mine waters were dominated by SO4, and elevated SO4 concentrations were typical also for mine waters from the two other areas. The increase in SO4 content of the mine waters was accompanied by elevated Fe, and this was a consequence of pyrite dissolution (Nordstrom, 1982), the most common Fe-bearing sulfide mineral in mineralized rocks of the studied mine sites. The acidity produced by dissolution of pyrite and other sulfides is readily neutralized by abundant Ca- and Mg-bearing carbonates as the waters were mostly neutral with some slightly acidic (Table 2). The molar Ca/Mg ratios of several mine waters of Dúbrava, Poprocˇ and Cˇucˇma ranged from 1:1 to 2:3. Such molar Ca/Mg ratios are characteristic of the dissolution of dolomites (Appelo et al., 1984) and other Mg-bearing carbonates such as magnesite. These carbonates are common in the mineralized rocks of the study areas. Waters of the main drainage streams below the mines had variable hydrogeochemistry and were clearly affected by discharge waters from adits and impoundments (Table 2). Downstream waters of Pernek and Poprocˇ maintained the same major ion chemistry (Ca–Mg–SO4) as the mine waters. Surface waters of the other three mine sites were less affected by mine discharges as their major ion chemistry was similar to that of upstream waters. This is due to dilution of the downstream waters

Fig. 9. Relationship between the As and Sb concentrations in leachates and their concentrations in mine wastes, stream sediments and soils.

612

E. Hiller et al. / Applied Geochemistry 27 (2012) 598–614

In general, the stream waters and groundwater from boreholes and household wells exhibited much higher concentrations of Sb than As (Fig. 7) and typically, Sb/As ratios of the waters ranged from 1.0 to 126. This is mainly due to fact that Sb is the primary component of the mineralized rocks as well as the mine tailings and waste-rocks, although differences in the mobility between As and Sb might be also responsible for these observations. Most of the mine waters discharging from adits and impoundments reflected the composition of the solids as indicated by the Sb/As ratio of >1.0 and up to 150. However, as one may see from the data presented in Fig. 7 and Table 2, there were some differences in the behavior between As and Sb in the aquatic environment. Impoundment discharge waters and groundwaters had somewhat higher Sb/As ratios than the mine tailings from which they were derived. Moreover, an increase of the Sb/As ratios in stream water samples collected farthest from mines at Dúbrava and Poprocˇ might be evidence of the different behavior between As and Sb. Although these observations could not be explained fully from the obtained results, some authors have suggested that such distribution changes between As and Sb result from either the higher affinity of As than Sb for the solid phases such as stream sediments and HFOs (Casiot et al., 2007; Masson et al., 2009; Milham and Craw, 2009) or the greater desorption of Sb than As from the solids (Ashley et al., 2003; Casiot et al., 2007). Data on the As and Sb contents of hydrous ferric oxides (Table 3) showed that the possible greater affinity of As than Sb for HFOs was relevant. These HFOs mainly consisted of ferrihydrite and goethite and contained much more As than Sb, excepting those from Pernek. Despite the lower Sb contents, HFOs at the study mine sites efficiently adsorbed both metalloids as confirmed previously by other studies (Belzile et al.,

2001; Roddick-Lanzilotta et al., 2002; Leuz et al., 2006). To test whether the susceptibility of Sb for leaching exceeded that of As, water leaching experiments were conducted on samples of mine wastes, stream sediments and soils. 5.3. Leaching tests Results showed that increasing amounts of As and Sb in the solids, resulted in increasing concentrations of both metalloids in leachate (Table 5 and Fig. 9). The concentrations of As and Sb in leachates from mine wastes were similar to those found in natural waters which were in contact with mine residues, although there were some notable differences. The Sb/As ratio of the leachates derived from mine residues were commonly higher than those of the solids, and this was in agreement with the field observations of higher Sb/As ratios in impoundment discharge and groundwaters than in tailings. However, comparing As and Sb concentrations of the leachates to those of the impoundment waters, most of the leachates contained more Sb than impoundment waters and this held also for As, although exceptions were observed. Leaching of both metalloids from mine residues under the natural conditions is a complex function of opposing processes which may lead to their attenuation or enhancement in waters. The most important attenuation processes considered in the context of this study were adsorption and precipitation reactions of As and Sb with secondary Fe oxides in the mine residues and HFOs deposited from discharge ˇ ucˇma, impoundment waters contained waters. At Poprocˇ and C more As than the leachates. Changes in redox conditions from oxic to anoxic within the impoundments might be responsible for the difference. Casiot et al. (2007) pointed out that As displayed higher

Table 6 Comparison of the As and Sb concentrations in waters (in lg/L) and stream sediments (in mg/kg) at the studied mine sites to some other mines located in Slovakia and worldwide. Location

Blackwater Au minea,b Salanfe (Switzerland)c Goesdorf Sb mine (Luxembourg)d Zlatá Idka Au–Sb mine (Slovakia)e Pezinok Sb mine (Slovakia)f Hillgrove Sb mine (Australia)g Sb mine of Bournac (France)h Baccu Locci Pb–As mine (Sardinia, Italy)i Endeavour Inlet (New Zealand)j Endeavour Inlet (New Zealand)k Türkönü Hg mine (Turkey)l Halıköy Hg mine (Turkey)m Reefton gold mine (New Zealand)n Macraes gold mine (New Zealand)n This study Pernek Sb mine Dúbrava Sb mine Medzibrod Sb mine Poprocˇ Sb mine ˇ ucˇma Sb mine C a b c d e f g h i j k l m n

Haffert and Craw (2008a). Haffert and Craw (2008b). Pfeifer et al. (2007). Filella et al. (2009). Rapant et al. (2006). Majzlan et al. (2007). Ashley et al. (2003). Casiot et al. (2007). Frau and Ardau (2003). Wilson et al. (2004a). Wilson et al. (2004b). Gemici and Tarcan (2007). Gemici et al. (2009). Milham and Craw (2009).

As

Sb

Water

Sediment

5–52,000 0.8–3922

47–258,000

0.5–104 11–80,900 13–7200 40–78 4.0–500 5.5–8.6 5.0–757 77–53,293 0.7–180 24–50,600 100–3000

4.9–974 39–147 90–1490 3000 1800–4800 14–190 28–5900

<1.0–5 <1.0–62 7.0–255 1.0–2150 1.0–1350

45–390 25–140 30–9180 52–5560 52–187

17–5900

Water

Sediment 10.3–426

11.5–2200 0.2–143 23–2500 470–55,000 2.1–32

10.6–1392 1.3–4880 11–317 80–4640 500

14–30 22–169 580–418,600 0.1–200 <100–1000 <1.0–100

18–243 34–1750

<1.0–31 4.0–9300 11–1290 5.0–1000 1.0–3540

48–703 46–644 32–890 215–1360 11–600

4.4–>2000

E. Hiller et al. / Applied Geochemistry 27 (2012) 598–614

mobility than Sb under simulated anoxic conditions, probably due to bacterial activity favouring the release of As. For sediments and soils, the Sb/As ratio was higher in most of the leachates than in the solids (Table 5), indicating again a higher mobility of Sb compared with As under oxidizing conditions. This result seems to agree well with lower As concentrations in the groundwaters extracted from boreholes in alluvial sediments as well as from household wells and increased Sb concentrations (Table 2) being several times above the drinking water limit of 5 lg/L (WHO, 2006). In oxidizing groundwaters such as those of the studied mine sites, Sb is expected to strongly sorb on clay minerals and Fe or Al oxyhydroxides (Leuz et al., 2006; Rakshit et al., 2011; Xi et al., 2011), thereby buffering Sb at low concentrations in these groundwaters. However, the combination of oxidizing conditions and the circumneutral pH of the groundwaters indicate that desorption from solid surfaces might also be important. Previous studies have revealed that Sb adsorption is greatest at low pH and decreases significantly with increasing pH (Leuz et al., 2006; Martínez-Lladó et al., 2008). Moreover, Martínez-Lladó et al. (2011) demonstrated that Sb retention in calcareous soils was low and not retarded during transport through soil columns. Thus, once Sb entered the groundwaters, its re-adsorption onto aquifer material surfaces would likely be minimal in these circum–neutral, oxidizing groundwaters. As a consequence, Sb seems to be relatively mobile in groundwaters at the studied mine sites. 5.4. Comparison with other mine sites To provide a general overview on contamination of the studied mine sites, the concentrations of As and Sb in waters and stream sediments collected in this study were compared to those previously reported from mine sites contaminated with As and Sb (Table 6). Studies of stibnite-dominant Sb mines in Luxembourg (Filella et al., 2009), France (Casiot et al., 2007), Australia (Ashley et al., 2003) and New Zealand (Wilson et al., 2004a,b), as well as in Slovakia (Majzlan et al., 2007; Rapant et al., 2006) all indicate highly elevated concentrations of As and Sb in stream sediments collected in the wider vicinity of these mines, reaching up to 5900 mg/kg As and 4880 mg/kg Sb. Published ranges of the As and Sb concentrations in stream sediments are clearly comparable to those measured in this study. Contamination of stream sediments with As and Sb is not restricted to the Sb mines with stibnite as the main ore mineral, but can be observed at other mine sites, e.g. cinnabar-dominant Hg mines. For example, stream sediments collected at the Halıköy Hg mine site are contaminated not only with Hg, but also contain significantly elevated concentrations of As and Sb, ranging from 17 to 5900 mg/kg and from 4.4 to >2000 mg/kg, respectively (Gemici et al., 2009). Published concentrations of As and Sb in waters affected by mining activities are approximately within the range of those determined in this study. The exceptions are a few notably high concentrations given in some studies such as Ashley et al. (2003) who reported up to 55,000 lg/L Sb in mine water collected at the discharge of the tailing dam at Hillgrove (Australia) or Haffert and Craw (2008a) who measured 52,000 lg/L As in wetland water of the Prohibition Mill site (New Zealand). These high concentrations of As and Sb in mine waters were likely due to dissolution of relatively soluble minerals such as arsenolite (As2O3) and valentinite (Sb2O3). 6. Conclusions Previous Sb mining at five sites in Slovakia (Pernek, Dúbrava, ˇ ucˇma) continue to contaminate the surMedzibrod, Poprocˇ and C rounding environment with As and Sb. Surface water samples collected from the mine sites showed variable levels of Sb (1–9300 lg/ L) as well as As (<1–2150 lg/L), dependent on the mineralogy of par-

613

ticular mine sites. Despite the natural attenuation processes (e.g. adsorption onto HFOs and dillution), concentrations of Sb in downstream surface waters and groundwaters remained high and were many times above the drinking water limit of 5 lg/L. Waters from each mine site showed no acid drainage (pH 6.2–8.2) because they are neutralized by abundant Ca- and Mg-carbonate intercalations in the mineralized rocks. The stream sediments contained up to 9180 mg/kg As and 1360 mg/kg Sb and concentrations of both metalloids decreased with distance from the mines. Sediments with the highest concentrations of As and Sb were in the vicinity of adit discharge waters. Compared to global average soil concentrations, As and Sb in most of the soil samples were highly elevated, reaching as much as 10,250 mg/kg and 9619 mg/kg, respectively. Hydrous ferric oxides commonly occurring at the studied mines consisted of ferrihydrite and goethite. As analyses revealed, substantial amounts of As and Sb were associated with these oxides. The mine wastes contained Fe oxides with variable concentrations of As and Sb, the most abundant products of the oxidative weathering of pyrite, arsenopyrite and stibnite contained in the mineralized rock complexes. These mine wastes exhibited elevated contents of As (5166 mg/kg) and Sb (9861 mg/kg). Water leaching experiments conducted with mine wastes, stream sediments and soils showed generally that under oxidizing conditions more Sb went into the water than As. This result was in agreement with the field observations of the predominance of Sb over As in most of the surface water and groundwater samples. Acknowledgments This study was financially supported by the Slovak Research and Development Agency under the Contracts Nos. APVV-026806 and APVV-VVCE-0033-07 ‘‘Scientific and Educational Centre of Excellence for Solid Phase Research Focused on Nanomaterials, Environmental Mineralogy and Material Technology (Centrum of excellence APVV-SOLIPHA)’’. References Alvarez, R., Ordóñez, A., Loredo, J., 2006. Geochemical assessment of an arsenic mine adjacent to a water reservoir (León, Spain). Environ. Geol. 50, 873–884. Anon, 1998. Directive of Ministry of Environment of the Slovak Republic No. 549/98 for the Risk Assessment Resulting from Contaminated Stream and Bed Sediments. Appelo, C.A.J., Beekman, H.E., Oosterbaan, A.W.A., 1984. Hydrochemistry of springs from dolomite reefs in the southern Alps of northern Italy. In: Eriksson, E. (Ed.), Hydrochemical Balances of Freshwater Systems. IAHS Publ. 150, pp. 125–138. Ashley, P.M., Craw, D., Graham, B.P., Chappell, D.A., 2003. Environmental mobility of antimony around mesothermal stibnite deposits, New South Wales, Australia and southern New Zealand. J. Geochem. Explor. 77, 1–14. Ashley, P.M., Craw, D., Tighe, M.K., Wilson, N.J., 2006. Magnitudes, spatial scales and processes of environmental antimony mobility from orogenic gold-antimony mineral deposits, Australasia. Environ. Geol. 51, 499–507. Baroni, F., Boscagli, A., Protano, G., Riccobono, F., 2000. Antimony accumulation in Achillea ageratum, Plantago lanceolata and Silene vulgaris growing in an old Sb mining area. Environ. Pollut. 109, 347–352. Belzile, N., Chen, Y., Wang, Z., 2001. Oxidation of antimony (III) by amorphous iron and manganese oxyhydroxides. Chem. Geol. 174, 379–387. Cambel, B., Khun, M., 1983. Geochemical characteristics of black shales from orebearing complex of the Malé Karpaty. Geol. Carpath. 34, 15–44. Casiot, C., Ujevic, M., Munoz, M., Seidel, J.L., Elbaz-Poulichet, F., 2007. Antimony and arsenic mobility in a creek draining an antimony mine abandoned 85 years ago (upper Orb basin, France). Appl. Geochem. 22, 788–798. Chovan, M., Rojkovicˇ, I., Andráš, P., Hanas, P., 1992. Ore mineralization of the Malé Karpaty Mts. (Western Carpathians, Slovakia). Geol. Carpath. 43, 275–286. Chovan, M., Háber, M., Jelenˇ, S., Rojkovicˇ, I., 1994. Ore Textures in the Western Carpathians. Slovak Academic Press, Bratislava. Chovan, M., Majzlan, J., Ragan, M., Siman, P., Krištín, J., 1998. Pb–Sb and Pb–Sb–Bi sulfosalts and associated sulphides from Dúbrava antimony deposit, Nízke Tatry Mts. Acta Geol. Univ. Comen. 53, 37–49. Chovan, M., Lalinská, B., Šottník, P., Jurkovicˇ, Lˇ., Zˇenišová, Z., Flˇaková R., Krcˇmárˇ, D., Lintnerová, O., Hiller, E., Klimko, T., Jankulár, M., Hovoricˇ, R., Ondrejková, I., Lux, A., Vaculík, M., Michnˇová, J., Petrák, M., 2010. The assessment of the effects of mining activities on the environment around abandoned Sb deposits of Slovakia

614

E. Hiller et al. / Applied Geochemistry 27 (2012) 598–614

with remediation proposals. Final report of the project APVV-0268-06 (in Slovak). Council of the European Communities, 1976. Council Directive 76/464/EEC of 4 May 1976 on pollution caused by certain dangerous substances discharged into the aquatic environment of the community. Off. J. L 129, 23–29. Craw, D., Wilson, N.J., Ashley, P.M., 2004. Geochemical controls on the environmental mobility of Sb and As at mesothermal antimony and gold deposits. Trans. Inst. Min. Metall., Appl. Earth Sci. 113, B3–B10. ˇ Curlík, J., Šefcˇík, P., 1999. Geochemical Atlas of the Slovak Republic, Part V, Soils. Soil Science and Conservation Research Institute, Bratislava (in Slovak). Filella, M., Belzile, N., Chen, Y.W., 2002. Antimony in the environment: a review focused on natural waters. I. Occurrence. Earth Sci. Rev. 57, 125–176. Filella, M., Philippo, S., Belzile, N., Chen, Y., Quentel, F., 2009. Natural attenuation processes applying to antimony: a study in the abandoned antimony mine in Goesdorf, Luxembourg. Sci. Total Environ. 407, 6205–6216. ˇ Flaková, R., Zˇenišová, Z., Drozdová, Z., Milovská, S., 2005. Distribution of arsenic in surface and groundwater in Kolársky vrch mining area (Malé Karpaty Mts.). Podzem. Voda 9, 90–103 (in Slovak with English summary). Frau, F., Ardau, C., 2003. Geochemical controls on arsenic distribution in the Baccu Locci stream catchment (Sardinia, Italy) affected by past mining. Appl. Geochem. 18, 1373–1386. Gebel, T., 1997. Arsenic and antimony: comparative approach on mechanistic toxicology. Chem. Biol. Interact. 107, 131–144. Gebel, T.W., 1999. Arsenic and drinking water contamination. Science 283, 1455. Gemici, Ü., Tarcan, G., 2007. Assessment of the pollutants in farming soils and waters around untreated abandoned Türkönü mercury mine (Turkey). Bull. Environ. Contam. Toxicol. 79, 20–24. Gemici, Ü., Tarcan, G., Somay, A.M., Akar, T., 2009. Factors controlling the element distribution in farming soils and water around the abandoned Halıköy mercury mine (Beydag˘, Turkey). Appl. Geochem. 24, 1908–1917. Haffert, L., Craw, D., 2008a. Mineralogical controls on environmental mobility of arsenic from historic mine processing residues, New Zealand. Appl. Geochem. 23, 1467–1483. Haffert, L., Craw, D., 2008b. Processes of attenuation of dissolved arsenic downstream from historic gold mine sites, New Zealand. Sci. Total Environ. 405, 286–300. Jankulár, M., Klimko, T., Jurkovicˇ, Lˇ., Lalinská, B., Šottník, P., Lintnerová, O., Šutriepka, M., 2008. Examination of abandoned Sb deposits by mineralogical methods in Poprocˇ (Slovakia). Chem. Listy 102, 380–382. Klimko, T., Chovan, M., Huraiová, M., 2009. Hydrothermal mineralization in stibnite veins of the Spiš-Gemer rudohorie Mts. Miner. Slov. 41, 115–132 (in Slovak with English abstract and summary). Kodeˇra, M., Andrusovová, G., Belešová, O., Briatková, D., Dávidová, Š., Fejdiová, V., Hurai, V., Chovan, M., Nelišerová, E., Zˇeniš, P., 1990. Topographic Mineralogy of Slovakia – Part 2, first ed. Slovak Academy of Sciences, Bratislava (in Slovak). Lalinská, B., Chovan, M., 2006. Hydrothermal mineralization in the Medzibrod deposit and the Sopotnická dolina valley. Miner. Slov. 38, 261–272. Landscape Atlas of the Slovak Republic, 2002. Ministry of Environment of the Slovak Republic and Slovak Environmental Agency, first ed. Bratislava. Leuz, A.-K., Monch, H., Johnson, C.A., 2006. Sorption of Sb (III) and Sb (V) to goethite: influence on Sb (III) oxidation and mobilization. Environ. Sci. Technol. 40, 7277– 7282. Majzlan, J., Lalinská, B., Chovan, M., Jurkovicˇ, Lˇ., Milovská, S., Göttlicher, J., 2007. The formation, structure, and ageing of As-rich hydrous ferric oxide at the abandoned Sb deposit Pezinok (Slovakia). Geochim. Cosmochim. Acta 71, 4206–4220. Martínez-Lladó, X., de Pablo, J., Giménez, J., Ayora, C., Martí, V., Rovira, M., 2008. Sorption of antimony(V) onto synthetic goethite in carbonate medium. Solvent extract. Ion Exchange 26, 289–300.

Martínez-Lladó, X., Valderrama, C., Rovira, M., Martí, V., Giménez, J., de Pablo, J., 2011. Sorption and mobility of Sb(V) in calcareous soils of Catalonia (NE Spain): batch and column experiments. Geoderma 160, 468–476. Maruška, M., Chovan, M., Ševc, J., 2000. Mineralogical and environmental evaluation of the settling ponds at the deposit Dúbrava in Nízke Tatry Mts. Slovak Geol. Mag. 6, 61–69. Masson, M., Schäfer, J., Blanc, G., Dabrin, A., Castelle, S., Lavaux, G., 2009. Behavior of arsenic and antimony in the surface freshwater reaches of a highly turbid estuary, the Gironde Estuary, France. Appl. Geochem. 24, 1747–1756. Michálek, J., Chovan, M., 1998. Structural-geological and mineralogical evaluation of the Dúbrava Sb deposit. Miner. Slov. 30, 25–35 (in Slovak with English summary). Milham, L., Craw, D., 2009. Antimony mobilization through two contrasting gold ore processing systems, New Zealand. Mine Water Environ. 28, 136–145. Nordstrom, D.K., 1982. Aqueous pyrite oxidation and the consequent formation of secondary iron minerals. In: Kittrick, J.A., Fanning, D.F., Hossner, L.R. (Eds.), Acid Sulfate Weathering. Soil Sci. Soc. Am. Spec. Publ. 10, pp. 37–56. Pfeifer, H.-R., Häussermann, A., Lavanchy, J.-C., Halter, W., 2007. Distribution and behavior of arsenic in soils and waters in the vicinity of the former gold-arsenic mine of Salanfe, Western Switzerland. J. Geochem. Explor. 93, 121–134. Pitter, P., 1999. Hydrogeochemistry. Institute of Chemical Technology, Prague (in Czech). Rakshit, S., Sarkar, D., Punamiya, P., Datta, R., 2011. Antimony sorption at gibbsite– water interface. Chemosphere 84, 480–483. Rapant, S., Dietzová, Z., Cicmanová, S., 2006. Environmental and health risk assessment in abandoned mining area, Zlata Idka, Slovakia. Environ. Geol. 51, 387–397. Reimann, C., de Caritat, P., 1998. Chemical Elements in the Environment: The Factsheets for the Geochemist and Environmental Scientist. Springer-Verlag, Berlin-Heidelberg. Reimann, C., Matschullat, J., Birke, M., Salminen, R., 2009. Arsenic distribution in the environment: the effects of scale. Appl. Geochem. 24, 1147–1167. Roddick-Lanzilotta, A.J., McQuillan, A.J., Craw, D., 2002. Infrared spectroscopic characterisation of arsenate(V) ion adsorption from mine waters, Macraes Mine, New Zealand. Appl. Geochem. 17, 445–454. Smedley, P.L., Kinniburgh, D.G., 2002. A review of the source, behaviour and distribution of arsenic in natural waters. Appl. Geochem. 17, 517–568. Tighe, M., Ashley, P., Lockwood, P., Wilson, S., 2005. Soil, water, and pasture enrichment of antimony and arsenic within a coastal floodplain system. Sci. Total Environ. 347, 175–186. Trtíková, S., Madejová, J., Kušnierová, M., Chovan, M., 1999. Precipitation and chemical composition of iron ochres in the pyrite and stibnite deposits in the Malé Karpaty Mts. Slovak Geol. Mag. 5, 179–186. Veselsky´, J., Forgácˇ, J., Mejeed, S.Y., 1996. Contamination of the soil and stream sediments in the Malé Karpaty Mts. Miner. Slov. 28, 209–218. Vrana, K., Rapant, S., Bodiš, D., Marsina, K., Lexa, J., Pramuka, S., Manˇkovská, B., Cˇurlík, J., Šefcˇík, P., Vojtaš, J., Daniel, J., Lucˇiviansky´, L., 1997. Geochemical atlas of Slovak Republic at a scale 1:1,000,000. J. Geochem. Explor. 60, 7–37. WHO, 2006. Guidelines for Drinking-Water Quality: Incorporating First and Second Addenda to Third Edition. World Health Organization, Geneva. (accessed 24.06.11). Wilson, N.J., Craw, D., Hunter, K.A., 2004a. Antimony distribution and environmental mobility at an historic antimony smelter site, New Zealand. Environ. Pollut. 129, 257–266. Wilson, N.J., Craw, D., Hunter, K.A., 2004b. Contributions of discharges from a historic antimony mine to metalloid content of river waters, Marlborough, New Zealand. J. Geochem. Explor. 84, 127–139. Xi, J., He, M., Lin, C., 2011. Adsorption of antimony(III) and antimony(V) on bentonite: kinetics, thermodynamics and anion competition. Microchem. J. 97, 85–91.