Natural attenuation of arsenic in soils near a highly contaminated historical mine waste dump

Science of the Total Environment 414 (2012) 546–555

Contents lists available at SciVerse ScienceDirect

Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Natural attenuation of arsenic in soils near a highly contaminated historical mine waste dump Petr Drahota a,⁎, Michal Filippi b, Vojtěch Ettler a, Jan Rohovec b, Martin Mihaljevič a, Ondřej Šebek c a b c

Institute of Geochemistry, Mineralogy and Mineral Resources, Charles University, Albertov 6, 128 43 Prague 2, Czech Republic Institute of Geology, Academy of Sciences of the Czech Republic, v.v.i., Rozvojová 269, 165 00 Prague 6-Lysolaje, Czech Republic Laboratories of Geological Institutes, Charles University, Albertov 6, 128 43 Prague 2, Czech Republic

a r t i c l e

i n f o

Article history: Received 22 July 2011 Received in revised form 1 November 2011 Accepted 1 November 2011 Available online 30 November 2011 Keywords: Arsenic Contaminated soil Historical mine waste dump Long-term attenuation

a b s t r a c t Arsenic-contaminated soils near historical As-rich mine waste in Jáchymov (Czech Rep.), resulting from the smelting and seepage of the mine waste pore water, were studied to examine As partitioning between solid phases and pore waters. Mineralogical and geochemical analyses showed that As is exclusively associated with unidentified amorphous Fe oxyhydroxides, poorly crystalline goethite and hematite as adsorbed and coprecipitated species (with up to 3.2 wt.% As). Adsorption of As by Fe oxyhydroxides is likely to be a major control on the migration of As in the soil pore water containing only up to 15 μg L− 1 As(V). The slight variations in the dissolved As(V) concentrations do not follow the total contents of As in the soil or adsorbed As, but appeared to be a function of pH-dependent sorption onto Fe oxyhydroxides. The geochemical modelling using PHREEQC-2 supported the efficiency of As(V) adsorption by Fe oxyhydroxides in the soil affected by As-rich waste solution seepage. It also suggested that active Fe oxyhydroxides has a strong attenuation capacity in soil that could effectively trap the aqueous As(V) from the unremitting waste seepage for the next approx. 11600 years. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Arsenic is a toxic metalloid that can be released into the environment through natural processes (rock weathering or volcanism) and from anthropogenic sources (mining, agricultural and industrial activities) (Matschullat, 2000; Smedley and Kinniburgh, 2002; Morin and Calas, 2006). Background As concentrations in soils do not generally exceed 40 mg kg− 1 (Smith et al., 1998), but can reach several thousands of mg kg− 1, usually in the surface horizons of soil contaminated by industrial, agricultural or mining activities (Smith et al., 1998; Smedley and Kinniburgh, 2002). More specifically, numerous mines exploited As-bearing ores in European countries as early as in the Roman and Middle ages (Shotyk et al., 1996; West et al., 1997; Costagliola et al., 2008). After extraction of economic elements, Asrich residual wastes were left unsecured in the close vicinity of the mines, inducing changes in the landscape and contamination of the surrounding ecosystems. When these wastes react with surface and rain water, As is usually transferred to the dissolved fraction and mobilized into surface water and groundwater. On the other hand, natural attenuation processes can occur with dissolved As frequently trapped by precipitation of secondary As minerals, coprecipitation/sorption reaction with Fe, Mn and Al oxyhydroxides, carbonates, clay minerals or complexed by humic substances (Stollenwerk, 2003; Drahota and

⁎ Corresponding author. Tel.: + 420 221951498; fax: + 420 221951496. E-mail address: [email protected] (P. Drahota). 0048-9697/$ – see front matter © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.scitotenv.2011.11.003

Filippi, 2009; Tokoro et al., 2010). Consequently, assessment of the environmental impact of historical mining residues depends on identification and quantification of the various As species occurring in the mine waste and surrounding soil profiles, and on the link between the evolution of the relative proportions of these species and the physicochemical characteristics of both the waste and the soil medium. This work presents the results of geochemical analysis and mineralogical characterization of pore water and soil from four soil profiles in the close vicinity of historical As-rich mine waste in the Jáchymov ore district (Czech Republic). This investigation was carried out to elucidate (1) how the As is sequestered in the soil, so that the long-term stability of As-bearing phases in the soil could be evaluated, and (2) how the interaction of As-rich waste solutions with the soil contributes to chemical evolution of the soil pore waters, with particular emphasis on As geochemistry. The role of the As mobility in mine waste is not examined in our study, since this was simultaneously studied and the results are given in the work by Filippi et al. (in prep). It is hoped that this study, although based on the specific site system, will also be of benefit to researchers working on As in soil systems world-wide. 2. Experimental methods 2.1. Site description and sample collection and preparation The Giftkies arsenic deposit is located in the District of Jáchymov in the north-western Czech Republic (Fig. 1a). Local mineralization comprises mica schist zones intensively greisenized and mineralized

P. Drahota et al. / Science of the Total Environment 414 (2012) 546–555

547

Ski tow GERMANY

POLAND

100 m 0 90

Jáchymov Prague

CZECH REPUBLIC

N

S L OVA K I A Bratislava Vienna Budapest

AU S T R I A

H U N G A RY

85 0 Probable position of the historical Ölbecken smelter (remnants unpreserved)

PD2

Hut PS1 Preserved remnants of the former

mining

o f As

1000

950

900

PS2

Legend: soil profile soil profile and lysimeter collapsed adit entrance selected contours

PS3

artificial sinkholes

a

waste dumps

Profile PD2

Profile PS1

Profile PS2

Profile PS3

buried soil beneath the waste dump

soil 10 m below the waste dump

60 m at the level of the dump

300 m from the waste dump in the undisturbed area

was te P D2 -4 PD2 (50-65 ): -5 ( 65PD2 88) : -6 ( 88120 ):

b

dum

p su rfac e

was te m ate form er s rial urfa dun ce soil

inte nsiv ely d un r ub bly

PS1 -1 PS1 (0-8): -2 ( PS1 8-24 ): -3 PS1 (24-3 0) -4 PS1 (23-5 : 0) -5 PS1 (50-8 : 5) PS1 6 (85-1 : 05 -7 PS1 (105- ): 1 -8 ( 130 30): -14 0) :

surf a

ce

fore st dar litter kb bro rown wn hu rust humu mus s oil y s rust clayey soil -san y cl a d y yell ow ey-san y soil sa d yell ow ndy so y soil s il w and rust ith y y-ye llow soil wit h wea ther ed

PS2 -1 PS2 (0-4): -2 PS2 (4-14 ): -3 PS2 (14-2 0) -4 PS2 (20-4 : 0) -5 PS2 (40-6 : 8) : -6 ( 6 PS2 8 -7 ( -105): 105 -13 0):

surface

surf a

ce

fore s dar t litter kb bro rown h wn u rust humu mus so s y il rust clayey- soil y s gray clayey andy s oil - sa ish gray -yello ndy so w il ishyell soil wi ow wea th ther ed

PS3-1 (0-5): PS3-2 (5-13): PS3-3a (13-28): PS3-3b (28-55): PS3-3c (55-75):

forest litter dark brown humus soil rusty-brown clayey soil rusty-brown clayey soil brown clayey soil

Fig. 1. (a) A simplified map of the historical mining remains of the Giftkies mine in the Jáchymov ore district. Location of the Ölbecken smelter is estimated from the map of Putz from 1783 presented in Bufka and Velebil (2005). (b) Descriptions of the horizons in the soil profiles (depth in cm).

by arsenopyrite. Exploitation of arsenopyrite-rich ore at the Giftkies arsenic deposit was probably concomitant with the smelting period of the nearby Ölbecken smelter (Fig. 1a) that occurred between

1618–1771 (Bufka and Velebil, 2005). The mining activity in the Giftkies mine at that time has resulted in simultaneous deposition of several waste dumps with extremely high As concentrations (up

548

P. Drahota et al. / Science of the Total Environment 414 (2012) 546–555

to 10 wt.% As). The primary arsenopyrite (FeAsS) has mostly been oxidized and scorodite (FeAsO4·2H2O), kaňkite (FeAsO4·3.5H2O) and pitticite (Fex(AsO4)y(SO4)z·nH2O) are currently the major As carriers in the dump. The minor As-bearing minerals in the waste are represented by arsenopyrite, Fe oxyhydroxides and arsenian jarosite (KFe3[(S,As)O4)](OH)6) (Filippi et al., 2009, in preparation). The waste dumps are located in a V-shaped valley of the Veseřice Creek on a mountain with a 25° slope (~920 m a.s.l.). Despite the long time of deposition and steep slope of the area, the mechanical erosion of the dump is negligible. This is probably due to afforestation (approx. 80% Norway spruce) of the area and natural revegetation (grasses) in the thin soil layer on the waste dump. The soil type of the area is Dystric Cambisol. The site has annual rainfall of approximately 760 mm and a mean annual temperature of 6 °C (Erbanová et al., 2008). Further details regarding the historical background and the geological setting of this site can be found in Ondruš et al. (2003) and Bufka and Velebil (2005). Soil/waste samples from four profiles were collected in June and November 2008. Their locations were selected to investigate the potential transport of As into soil from the vertical seepage of solutions percolating through the waste dump. The PD2 site represents a buried soil profile located at the bottom of the dump and can be thus influenced by direct seepage of As-rich waste solutions; profile PS1 is located 10 metres from the dump down the valley and can be mainly influenced by erosion of waste material or intermittent flow during rainfall events. Profiles PS2 and PS3 in a wider vicinity of the waste dump represent sites without distinct influence of the studied waste dump. Profile PS2 is located at the same level as the waste dump approximately 60 metres south of the dump and the surface around reveals some historical anthropogenic impact. Profile PS3 is located approximately 300 m southwest of the dump in the local plateau and represented an area undisturbed by mining activities. The description of the soil profiles is given in the Fig. 1b. In the field, each soil sample was collected and homogenised in a polypropylene bag for soil pH and major and trace element analyses. In addition, samples from different depths of each soil horizon were intended for mineralogical characterization. The bulk soil density in the PD2 profile was determined by core method which involves determining the mass and volume of the soil sample. Soil solutions were sampled in the field repeatedly between June 2009 and November 2010 using suction lysimeters (SPE20, UMS) with a double-0.2 μm nylon membrane (Wenzel et al., 1997). The lysimeters were placed immediately above the unsaturated C-horizon to monitor the soil solution leaving the developed soil profile. To install each lysimeter, a 5 cm soil auger was used to excavate a hole in the topsoil very close to the soil profiles PD2, PS1 and PS2 (Fig. 1a). The excavated soil was placed on a polyethylene sheet; each horizon was sieved to obtain the b0.5 mm fraction, which was mixed to a wet slurry with distilled water. The slurries were subsequently placed in the hole around the lysimeter, and the area around the top of the lysimeter was filled with the remaining topsoil. Lysimeters were allowed to settle for one year, and then purged by vacuum pumping to a pressure of 0.7 bars. The temperature, pH, Eh, and specific conductivity of the pore water were measured in the field, after stabilization, with calibrated portable multimeters (WTW). The sensors were calibrated using NIST-traceable calibration (WTW technical buffers) immediately prior to the measurement. The accuracy of the pH, Eh, and specific conductivity measurements was ±0.01 std. pH, ±0.5 mV, and ±0.001 mScm − 1, respectively. The field-measured redox potential values were referred to the standard hydrogen electrode. Pore water samples were divided into precleaned HDPE bottles, and those destined for major cation and trace element analysis were acidified to pH b 2 by adding 2% HNO3 (suprapure, Merck). Samples for anionic species were left unpreserved. All the water samples were stored in a refrigerator until the analysis was completed.

2.2. Chemical and mineralogical analyses The soil pHH2O was measured within 8 h of sample collection using a 1:2.5 (w/v) ratio of untreated homogenised solid–water suspension after 1-hour agitation (Pansu and Gautheyrou, 2006). The concentrations of SiO2, Al2O3, CaO, MgO, Na2O, K2O, P2O5, SO3, S -2 + S 0, Mn, Fe, Ba, Zn, Cu, Pb and As in the b2 mm fraction of airdried and homogenised soils were measured using an X-ray fluorescence spectrometer (XRF, ARL 9400 XP +) with the Uniquant TM 4 analysis program (UNIQUANT 4, 1999). XRF Uniquant is a semiquantitative method and the associated errors were estimated to be ±0.2 wt.% for SiO2, ±0.1 wt.% for other major constituent oxides and ±0.01 wt.% for minor elements. The total organic carbon (TOC) was quantified by dry combustion with infrared detection using a ELTRA CS 530 total carbon analyzer and ELTRA CS 500 total inorganic carbon analyzer (Eltra, Germany). In order to characterize As-bearing minerals in the soils, heavyand light-minerals were separated using bromoform diluted with 1,4-dioxane (d = 2.81 g cm − 3) in the 0.1-0.315 mm fraction (according to previous studies of Filippi et al., 2007; Drahota et al., 2009) of 15 samples representing the upper and deeper mineral horizons in the soil profiles. Minerals were mounted on glass slides using polyester resin and thin sections were prepared. These polished thin sections of the heavy- and light-mineral fractions were examined by scanning electron microscopy (SEM) and electron microprobe analysis (EMPA). The backscattered images and compositional maps were obtained on a SEM (CamScan S4) equipped with an Oxford Link energy dispersive spectrometer (EDS) and a Link ISIS 300 microanalytical system. For EMPA, we used a Cameca SX 100 instrument equipped with four wavelength dispersion spectrometers (WDS), operating at 15 kV and 10 nA with 2 μm beam resolution and 10 s counting per element. The X-ray diffraction analyses (XRD) of the bulk soil samples, heavy and light mineral fractions and separated particles from the heavy mineral fraction were performed using a PANalytical X'Pert Pro diffractometer equipped with a diffracted-beam monochromator and X'Celerator multichannel detector. The analysis conditions were as follows: CuKα radiation, 40 kV, 30 mA, step scanning at 0.02°/250 s in the range 3–70° 2θ. The qualitative analysis employed the X'Pert HighScore software 1.0 d, equipped with the JCPDS PDF-2 database (ICDD, 2002). The major cations in pore water samples were analysed by ICPOES (IRIS Intrepid II XPS, USA). Quality control was performed by inserting a QC sample into the analytical run after each 10 samples. Field and laboratory duplicates indicate a relatively high level of reproducibility (b10%). Trace elements were determined by quadrupole ICP-MS (ThermoScientific XSeries II). The measured concentrations of trace elements in the certified and synthetic standards (NIST 1640, 1643 d) were within ±5% of their certified values. Arsenic(III) and As(V) were determined for the first five pore water samples using the HG-CT-ICP-OES method slightly modified from Richter et al. (2004) for which a CO2 trap is in the experimental setup was added. Because As(III) was below the detection limit of 1.5 μg L − 1 in all the pore water samples, inorganic As speciation was not employed in samples from the subsequent sampling campaigns. Anions (SO4 2−, Cl −, NO3 −) were analysed by ion chromatography (HPLC, columns Dionex ICS-2000). The relative error of these analytical determinations was always b15%. 2.3. Selective chemical extractions In order to limit possible As re-adsorption and redistribution phenomena that frequently occur when employing sequential extractions (Van Herreweghe et al., 2003), three single chemical extractions were performed in parallel on bulk samples (b2 mm). A 0.05 M solution of (NH4)2SO4 (sulphate extraction) was used to remove “exchangeable As”, which is related to weakly sorbed As. A 0.05 M

P. Drahota et al. / Science of the Total Environment 414 (2012) 546–555

solution of (NH4)2HPO4 (phosphate extraction) was used to extract more strongly sorbed As (Keon et al., 2001; Wenzel et al., 2001; Cancès et al., 2008). For these extractions, the solid-to-solution ratio was set at 1 g of solid in 40 mL solution. A 0.2 M solution of ammonium oxalate/oxalic acid (NH4)2C2O4/H2C2O4, pH = 3 (oxalate extraction) was used under dark conditions to release very strongly associated As (by sorption or co-precipitation) with amorphous and poorly crystalline Fe oxyhydroxides and possibly reactive ferric arsenates by dissolution of these species (Jackson et al., 1986; Filippi et al., in prep). Because of the potentially high extractable Fe in this extraction, the solid-to-solution ratio was set at 0.4 g of solid in 40 mL solution to ensure that the extractant did not become exhausted (oxalate fully complexed by iron, leading to iron oxide dissolution being limited by the oxalate concentration) (Parfitt, 1989). The suspensions were gently agitated in a table shaker (60 rpm) at a temperature of 23 ± 2 °C. After each extraction, the solid samples were centrifuged at 3500 rpm for 5 min. The remaining solid materials were washed with 40 ml of deionised water (MilliQ-plus) and centrifuged at 3500 rpm for 5 min. After extraction and washing, supernatant fluids were mixed and passed through 0.2 μm membrane filters (Millipore®) using the Sartorius filtration device. Subsequently, the extracts were diluted with 2% HNO3 prior to further analysis of Al, As, Fe, Mn and S by ICP-OES. Every fifth extraction was performed in duplicate (five in total), exhibiting a standard deviation of less than 15%; procedural blanks were run during the extraction procedures. 2.4. Thermodynamic modelling The PHREEQC-2 geochemical speciation-solubility code, version 2.13.2 for Windows (Parkhurst and Appelo, 1999) was used to determine the degree of saturation of soil pore water with respect to the As mineral phases. The thermodynamic database was supplemented by the solubility products of hydrous ferric arsenates (scorodite, pitticite) recently compiled by Drahota and Filippi (2009). Equilibrium geochemical modelling, including surface complexation reactions, was performed to examine the predicted effect of the change in concentrations of aqueous As and hazardous metals (Cu, Pb and Zn) in soil profiles with respect to their aqueous levels in the waste. The water chemistry data used in the inverse modelling are summarized in Appendix A and are represented by the average composition of pore water percolating the waste dump (Filippi et al., in preparation). The other data used in the model included the Fe oxyhydroxide surface characteristics, including the proportionality relationship, which is that 1/10 of each molecule of Fe(OH)3 is active as a sorbent, and that this Fe oxyhydroxide has a surface area of 5.28 × 10 4 m 2 mol − 1. The simulations were performed using PHREEQC-2 and run with the wateq4f.dat database, which includes surface complexation constants from Dzombak and Morel (1990). 2.5. Data processing The basic statistics and graphical interpretation of the data were performed by Excel 2003 (MS Office, Microsoft, USA) and SigmaPlot 11 (Systat Software, USA). The correlation coefficients were calculated using the Statistica software (StatSoft, USA). 3. Results 3.1. Arsenic distribution in soil Because exploitation of As-rich ore predominated at the Giftkies mine, attention was paid mainly to As within the studied profiles, despite the fact that Cu, Pb and Zn have also been found in elevated concentrations in the soil (b386 mg kg − 1 for Zn, b863 mg kg − 1 for Cu and b761 mg kg − 1 for Pb) (Table 1). The concentrations of

549

As are high at the upper horizon of the studied soil profiles (3700 mg kg − 1, 2330 mg kg − 1, 1870 mg kg − 1 and 210 mg kg − 1 for As at PD2, PS1, PS2 and PS3, respectively). They decrease rapidly downward to a few hundreds (PD2, PS1 and PS2) or tens (PS3) of milligrams per kilogram at the bottom of the soil profiles (Fig. 2, Table 1). In the top horizons, the As concentrations of highly contaminated profiles underlying the waste dump (PD2) and with distinct anthropogenic impact (PS1 and PS2) are approx. 10–20 times higher than the local As background (PS3); in the deep horizons, the As concentrations are less than 4 times higher than the local As background. These vertical distributions of As along the soil profiles are in agreement with contamination originating from superficial deposition, which imply that there is no migration downward the soil profile. The concentration of hazardous metals (Cu, Pb and Zn) in deep contaminated soil horizons (PD2, PS1, PS2) compared to their levels in the upper horizons may indicate different behaviour for the metals and As (Table 1). Their accumulation in deep horizons suggests a significant migration of these elements downward the soil profiles. In contrast, As concentrations in the deep horizons are significantly lower than in the upper horizons and approaching the local geochemical background, suggesting preferential accumulation of As in the upper horizons.

3.2. Arsenic associations with the solid phase The XRD analyses performed on the bulk soil samples, light and heavy mineral fractions and selected particles from heavy mineral fractions did not lead to identification of crystalline As minerals present in the waste, such as scorodite, kaňkite, arsenian jarosite and arsenopyrite. The mineralogy found in heavy mineral fractions was generally similar. For all depths and all profiles, poorly crystalline goethite (α-FeOOH) predominates in the XRD patterns of the >2.81 g cm − 3 fraction, with smaller or similar amounts of quartz, anatase, rutile, schorl, muscovite and chlorite. The broadening of the goethite reflections in all XRD patterns suggests poor crystallinity of the mineral (Appendix B). The minor ferric oxide in the PS1 and PS2 profiles is hematite (α-Fe2O3), which has not been detected in the PD2 profile. XRD did not suggest the presence of amorphous Fe oxyhydroxides either in heavy mineral fractions or in the individual loose grains. The absence of hydrous Fe arsenates, arsenian jarosite and arsenopyrite was confirmed by detailed examinations of heavy and light mineral fractions by microprobe and SEM. Surprisingly, these substances were not detected even in soil horizon PD2-4, which is located 17 cm under the waste material with abundant primary and secondary As minerals mentioned above. This observation is supported by high molar Fe/As values (>23) in the As-bearing compounds identified by microprobe (Table 2). Therefore, the only carriers of As identified are abundant goethite, hematite and possibly unidentified amorphous Fe oxyhydroxides. They were found in the form of coatings on the primary alumosilicates (Fig. 3a) and more frequently as individual loose grains with colloform textures and no relics of the primary minerals (Fig. 3b and c). The As contents in compact Fe oxyhydroxide grains usually vary on a microscale (Fig. 3d), suggesting different As concentrations in the parental solutions during Fe oxyhydroxide precipitation. Simple statistics of As contents in Fe oxyhydroxides also show the difference at the macroscale, between (i) the upper and deeper horizons within each soil profile and (ii) the different studied soil profiles (Table 2). The As contents in Fe oxyhydroxides are always approx. twice as high in the upper horizons than in the deeper horizons (Table 2). This was not observed in reference profile PS3 (Table 2), because the As content in Fe oxyhydroxides here is very low (b0.24 wt.% As) and generally close to the detection limit of the microprobe (dl ~ 0.04 wt.% As). These results are in good agreement with the distribution of total As in the soil

550

P. Drahota et al. / Science of the Total Environment 414 (2012) 546–555

Table 1 Selected XRF data for soil samples. Samples

Depth (cm)

pH

LOI (wt.%)

TOC (wt.%)

Al2O3 (wt.%)

Fe2O3 (wt.%)

S (mg kg− 1)

As (mg kg− 1)

Zn (mg kg− 1)

Cu (mg kg− 1)

Pb (mg kg− 1)

PD2-4 PD2-5 PD2-6 PS1-1 PS1-2 PS1-3 PS1-4 PS1-5 PS1-6 PS1-7 PS1-8 PS2-1 PS2-2 PS2-3 PS2-4 PS2-5 PS2-6 PS2-7 PS3-1 PS3-2 PS3-3a PS3-3b PS3-3c

50–65 65–88 88–120 0–8 8–24 24–30 30–50 50–85 85–105 105–130 130–140 0–4 4–14 14–20 20–40 40–68 68–105 105–130 0–5 5–13 13–28 28–55 55–75

3.84 3.96 4.27 5.12 4.18 4.13 4.25 4.63 4.80 4.52 4.46 4.56 3.72 3.76 4.27 4.52 4.41 4.44 4.91 3.86 4.19 4.31 4.28

nd nd nd 20.68 25.03 6.54 8.64 8.72 4.30 2.18 1.84 23.23 15.86 9.26 11.33 4.86 2.64 2.78 nd nd nd nd nd

nd nd nd 23.97 21.63 8.28 5.85 2.47 1.12 0.56 0.38 36.51 18.87 10.51 5.98 2.82 0.76 0.89 nd nd nd nd nd

21.45 22.88 24.49 18.57 19.23 20.09 22.63 23.33 24.33 21.75 22.62 14.59 19.54 21.03 25.60 24.02 22.59 22.39 17.26 19.52 22.13 21.67 21.62

5.22 11.40 6.76 8.36 7.39 7.12 7.89 7.66 5.96 5.43 5.12 8.05 6.45 10.51 8.98 7.78 6.63 5.74 8.96 8.30 8.70 8.38 8.65

861 1738 1410 2895 2903 713 685 1161 813 681 665 6808 2523 1410 1189 953 825 689 4889 1930 1049 1001 1269

3697 1063 352 1128 1428 2334 482 242 234 186 164 606 939 1865 495 482 365 220 207 151 76 74 80

195 267 386 290 239 176 165 280 299 266 228 345 249 231 264 362 320 199 278 174 198 239 239

711 863 423 168 184 94 100 141 154 140 121 214 145 171 218 243 230 145 22 35 b 20 b 20 b 20

761 613 236 455 566 317 238 235 186 179 148 724 585 724 371 310 218 124 322 209 b24 b24 88

Nd, not determined.

primarily in the oxalate extraction (targeting amorphous Fe oxyhydroxides and Al oxyhydroxides, as well as some portion of poorly crystalline Fe oxyhydroxides such as poorly crystalline goethite). We operationally defined this fraction to active Fe oxyhydroxides. In fact, this extraction removes between 17 and 65% of total As through the soil profiles, except for the organic-rich topsoil sample PS2-1 (0–4 cm) containing 36.5 wt.% TOC, where only 5% of total As is released. The largest amounts of extracted As and Fe occurred in the upper mineralized horizon, with extracted and total amounts of As and Fe generally decreasing with depth in the profiles. These results suggest a major association of As with active Fe oxyhydroxides that is also supported by strong correlations between As in the phosphate and oxalate fractions and total As (r = 0.94, p b 0.0001; r = 0.97, p b 0.0001, respectively). Any remaining As that has not been extracted by oxalate extraction could probably be attributed mainly to unaltered portions of crystalline Fe oxyhydroxides (mainly goethite), Fe oxides (hematite) and different substrates (e.g. organic matter in topsoil sample PS2-1). This suggestion is in agreement with XRD and SEM/microprobe data, which show a positive correlation between As in the oxalate fraction and the average

profiles (Fig. 2), suggesting that Fe oxyhydroxides could be almost the only As carrier in the studied soil samples.

3.3. Selective chemical extractions The results of single extractions of As from soil profiles are compared in Fig. 4 and Appendix C. All the profiles exhibit similar results for the single extractions employed. Very little As was removed in the sulphate extraction (0.12% to 0.89% of the total soil As), which targets readily exchangeable ions (outer-sphere complexes). This result suggests that As in soil is not readily mobile, but rather is tightly adsorbed onto mineral or organic surfaces or is present within insoluble mineral co-precipitates. Substantial amounts of As were extracted in the phosphate solution, which extracts strongly bound, adsorbed ions. Phosphate extraction released between 3.3 and 28.2% of the total As, depending on the depth (Fig. 4). In general, the amount of adsorbed As in the upper soil horizon is significantly greater (up to two orders of magnitude in the PS2 profile) than the amount of As adsorbed in the deeper horizons. Arsenic was extracted

Fe (g kg-1)

pH 3 3.5 4 4.5 5 5.5 6 0

0

20

40

60

As (mg kg-1) 80

0

1000 2000 3000 4000 5000

20

Depth (cm)

40 60

Waste Soil

80 100 120

Waste/soil profile PD2 Soil profile PS1 Soil profile PS2 Reference soil profile PS3

140 Fig. 2. Depth patterns of pH, total Fe and As in soil profiles (PS1, PS2, PS3) or waste/soil (PD2). Note that the waste/soil boundary is valid only for the PD2 profile.

P. Drahota et al. / Science of the Total Environment 414 (2012) 546–555

551

Table 2 Electron microprobe microanalyses of arsenic and iron (wt.%) in the heavy mineral fraction (d > 2.81 g cm− 3) of the upper soil horizons and deep soil horizons. Samples

Depth (cm)

n

Mean Fe2O3,tot (1 s)

Fe2O3,tot range

Mean As (1 s)

As range

Mean Fe/As

Fe/As range

PD2-4 PD2-6 PS1-2 PS1-8 PS2-2 PS2-7 PS3-2 PS3-3c

50–65 88–120 8–24 130–140 4–14 105–130 5–13 55–75

18 16 20 20 20 20 15 15

78.6 77.1 78.1 79.6 79.0 77.0 85.7 83.1

66.5–97.7 66.8–83.4 65.6–97.0 63.9–96.4 73.5–94.1 71.3–82.5 77.5–95.7 74.4–92.9

1.04 0.50 1.17 0.53 1.35 0.67 0.07 0.08

b 0.04–2.59 0.11–1.74 0.07–3.24 b 0.04–1.79 b 0.04–3.04 b 0.04–2.37 b 0.04–0.16 b 0.04–0.24

370 290 220 1010 260 430 2520 2050

>27 42–670 22–1220 >50 >23 >31 >450 >290

(8.9) (4.0) (8.2) (8.6) (6.0) (2.8) (6.5) (7.1)

(0.71) (0.43) (0.86) (0.39) (0.82) (0.56) (0.05) (0.06)

Notes: n is the number of analyses; Fe/As are molar ratios; oxidation state of iron for which multiple oxidation states are known, is assumed.

spot analyses of As in Fe oxyhydroxides by microprobe (apart from a PD2-4 sample outlier, the data show good correlation r = 0.92, p = 0.003). 3.4. As in the pore water Table 3 depicts the pore water concentrations for As and selected physico-chemical characteristics in soil at three lysimeter stations underlying the waste dump (PD2) and near the waste dump (PS1 and PS2) (Fig. 1a). Levels of pore water As(V) (based on speciation analyses of pore water) were not significantly different between the lysimeter stations at any time of sampling, with average concentrations of about 3.2 ± 2.2, 12.6 ± 1.4 and 4.1 ± 0.7 μg L − 1 for the PD2, PS1 and PS2 lysimeter stations, respectively. Surprisingly, dissolved As does not follow the total As in the soil even at the PD2 lysimeter station, which is located approx. 17 cm below the waste dump with solutions containing 2.5–3.5 mg L − 1 As (Filippi et al., in preparation). There is also no relationship between dissolved As and adsorbed As loads that were determined by considering the sorbent as the sum of Fe liberated by the oxalate extraction and adsorbed As, as determined from phosphate extraction. The slight variability in aqueous As between the stations could be attributed only to variations in the pore water pH (Table 3). The dissolved concentrations of other hazardous metals (Cu, Pb and Zn) are generally higher than that of As in spite of their lower total contents in the soil. The levels of Cu, Pb and Zn in the pore water were also significantly different between the lysimeter stations (Table 3). The lowest concentrations of the metals were found in the PS2 lysimeter station (b20 μg Cu L − 1, b1.8 μg Pb L − 1b, 500 μg

Zn L − 1), while concentrations of Zn usually exceeded 1000 μg L − 1 in PD2 and PS1, as did the concentrations of Cu in the PD2 lysimeter stations. Pore water concentrations of Pb were significantly lower than those of Cu and Zn in all the lysimeter stations. Different pore water-soil partitioning of As and metals point out the different mobilities of these elements in the soil system with decreasing mobility in the sequence Zn > Cu > Pb > As (based on partition coefficient between the solid and aqueous phases).

3.5. Geochemical modelling Geochemical models of the soil pore waters predict that the saturation indices (SI = log(IAP/Ksp)) for hydrous ferric arsenates (scorodite and pitticite) and jarosite are always lower than −0.3 for different Ksp (logKsp = −25.86 and −25.40 for scorodite according to Langmuir et al., 2006 and Bluteau and Demopoulos, 2007, respectively; logKsp = −23.00 and −24.14 for pitticite according to Langmuir et al., 2006 and Paktunc and Bruggeman, 2010, respectively). At relatively low Fe concentrations and the given pH values, the ferrihydrite (5Fe2O3∙9H2O) is undersaturated in pore water in the PD2 profile (SI b −0.4) while it is usually (−0.5 b SI b 0.2) and always (SI > 0.2) saturated in pore waters in the PS2 and PS1 profiles, respectively. However, a higher SI does not necessarily predict the occurrence of ferrihydrite (Ksp = 10 − 39), since it is metastable with respect to goethite (Ksp = 10 − 41) and hematite (Ksp = 10 − 43), which exhibit supersaturation in all the pore waters studied (SI > 3.0 for goethite and SI > 8.0 for hematite). The saturation indices for the As-bearing phases are given in the Appendix D.

Fig. 3. SEM images of Fe oxyhydroxides from the PS2 soil profile. (a) Backscattered image with phylosilicates (dark) and Fe oxyhydroxides coatings (light). (b, c) Backscattered images of loose Fe oxyhydroxides grains with zoned textures and corresponding X-ray maps for (d ) As, (e) Al, and (d) Si.

552

P. Drahota et al. / Science of the Total Environment 414 (2012) 546–555

Depth (cm)

0

PD2

PS1

PS2

PS3

% of total As

% of total As

% of total As

% of total As

20

40

60

80 100

0

20

40

60

0

80

20

40

60

0

0

0

0

20

20

20

20

40

40

40

60

60

60

80

80

80

80

100

100

100

120

120

120

140

140

40

Waste Soil

60

0

20

40

60

Sulphate Phosphate Oxalate

Fig. 4. Depth variations of selective chemical extractions of As applied to bulk soil horizons (b2 mm fraction) for the studied profiles.

The attenuation capacity of the soil underlying the waste dump could be estimated from the potentially unequilibrated level of Fe oxyhydroxides in the soil. The total content of active Fe oxyhydroxides in the soil that may actively adsorb As was estimated from the Fe concentration in the extraction of 0.2 M NH4–oxalate/oxalic acid (2 h in the dark) (Cornell and Schwertmann, 1996). Based on the ideal goethite and ferrihydrite formula (FeOOH ~ 88.858 g mol − 1, 5Fe2O3·9H2O ~ 192.087 g mol − 1), the estimated mass of active Fe oxyhydroxides in the soil underlying the waste dump (PD2 profile) would corresponded to approximately 36–39 g kg − 1. The proportion of Fe oxyhydroxide was multiplied by the bulk soil density (1.52 g cm − 3). The resulting content of active Fe oxyhydroxides in the upper mineralized horizons (0.4 m), which effectively immobilized As, would corresponded to approx. 55000 g m − 3 of poorly crystalline goethite or to approx. 60000 g m − 3 of ferrihydrite in the soil. This value is greatly in excess of the amount of equilibrated Fe oxyhydroxides inferred from the surface complexation modelling (1760 g m − 3) and suggests a high attenuation capacity of active Fe oxyhydroxides in the soil.

Equilibrium geochemical modelling was performed to examine the predicted effect of the decline of aqueous As(V) percolating vertically from the waste into the soil as a result of sorption on Fe oxyhydroxides. The adsorption simulation suggested equilibration of average waste pore water including a complex system of sorbents with different amounts of Fe oxyhydroxides. Assuming the adsorption of As(V) and other elements from the waste solution on active Fe oxyhydroxides as being the most important sorbent in the soil system, the adsorption simulation shows that 5–6 mg of Fe oxyhydroxides is sufficient for reducing the concentration of As(V) from 2651 μg L − 1 to the range 1.9–7.5 μg L − 1, as observed in the pore waters of the waste and underlying soil (PD2 lysimeter station), respectively (Fig. 5). The similar adsorption simulation shows that an analogous amount of Fe oxyhydroxides (2–6 mg L − 1 of waste solution) can also reduce the aqueous concentrations of Cu, Pb and Zn in waste to those found in soil pore water. As seen in Fig. 5, the adsorption efficiency of As(V) and metals under the observed conditions (2–6 mg of Fe oxyhydroxides per litre of waste solution) is sensitive to the pH and decreases in the sequence from As > Cu, Pb > Zn. To adapt the adsorption modelling results for As to natural conditions, the proportion of Fe oxyhydroxides (5–6 mg L − 1 of waste solution) from the adsorption simulation was multiplied by the annual seepage flux of water from the waste dump into the soil (334 L m − 2 yr − 1), which was estimated from the average annual precipitation (776 mm, Erbanová et al., 2008) and precipitation/ discharge ratio (0.44, Hruška and Krám, 2003; Oulehle et al., 2007) in the studied area. These quantifications show that As in the annual seepage flux (886 mg m − 2 yr − 1) would be sequestered in the upper soil horizons (0.4 m) by approximately 4.2–5.0 g m − 3 of Fe oxyhydroxides (i.e., assuming same surface area and site density for the Fe oxyhydroxides synthesized in the laboratory and in the field). Consequently, the model predicts that nearly 1760 g m − 3 of Fe oxyhydroxides has equilibrated with As and other elements in the soil underlying the waste, assuming similar annual input of As over a period of 350 years, representing the age of the waste dump.

4. Discussion 4.1. Mechanisms controlling the As mobility The distribution of As and several hazardous metals (Cu, Pb and Zn) in the soil profiles reflects the contamination originating from superficial deposition that was derived from surface mining activities and/or local smelting of As-rich ores at the smelter. The soil underlying the waste dump (PD2 profile) has been contaminated from the vertical seepage of solutions percolating through the waste dump. The similar group of elements contaminating the soil underlying the waste dump and its vicinity (PS1 and PS2 profiles) suggests a similar source of contamination. All the results suggest that, once released into the soil from the atmosphere during smelting and/or surface mining activities or

Table 3 Geochemistry of soil pore water samples from lysimeter stations at the Giftkies mine area. Lysimeter station

PD2 PS1 PS2

n

5 2 5

Mean pH (1σ)

4.60 (0.10) 5.69 (0.33) 5.09 (0.20)

pH range

4.51–4.58 5.45–5.92 4.89–5.41

Mean Fe (1σ)

Fe range

Mean As (1σ)

As range

Mean Cu (1σ)

Cu range

Mean Zn (1σ)

Zn range

Mean Pb (1σ)

Pb range

μg L− 1

μg L− 1

μg L− 1

μg L− 1

μg L− 1

μg L− 1

μg L− 1

μg L− 1

μg L− 1

μg L− 1

14.0 (7.8) 110.6 (18.0) 103.0 (19.6)

5.1–23.9 97.8–123.3 78.3–121.6

3.2 (2.2) 12.6 (1.4) 4.1 (0.7)

1.9–7.5 11.6–13.6 3.6–5.0

1136 (183) 40 (8) 20 (2)

905–1318 34–45 18–23

1040 (226) 1362 (326) 418 (147)

891–1321 1131–1592 200–506

10.5 (7.2) 28.6 (6.9) 1.8 (1.0)

1.7–15.9 23.7–33.4 0.8–2.8

10000 1000

7

pH

6

As Cu Pb Zn

100 5

pH

dissolved concentration (µgL-1)

P. Drahota et al. / Science of the Total Environment 414 (2012) 546–555

10 4

1 0.1 0.1

3 10

1

Fe oxyhydroxide equilibrated (mg

L-1)

Fig. 5. Surface complexation modelling results showing predicted dissolved As, Cu, Pb and Zn concentrations equilibrated with different amounts of Fe oxyhydroxides. The range of dissolved As concentrations observed in the soil pore water under the waste (PD2 profile) is shown as the shaded area. Note the different scales for dissolved concentrations and pH. The initial chemistry of the solution before equilibration with Fe oxyhydroxides corresponds to the average pore water percolating through the waste dump with 2651 μg L− 1 As(V), 1306 μg L− 1 Cu, 14 μg L− 1 Pb, 1126 μg L− 1 Zn and pH 3.765. The other input parameters included proportionality relationship 0.1 sites mole− 1 for Fe oxyhydroxides; specific surface 5.28 × 104 m2 mol− 1 of Fe oxyhydroxides. Database: wateq4f.dat.

from overlying waste, the As was readily scavenged by Fe oxyhydroxides. In contrast, the efficiency of sorption processes to reduce As migration cannot be suggested for superficial contamination by Cu, Pb and Zn (Fig. 5). Composition and extraction data show that As is associated with Fe oxyhydroxides as an adsorbed and coprecipitated species. The small amount of As removed in the sulphate extraction step (0.4% on average), attributed to weakly bound ions, suggests that only a negligible fraction was bound in outer-sphere complexes. More As (7.5% on average) was removed in the phosphate extraction, which is consistent with the interpretation that As is mostly a strongly bound complex (inner-sphere complex) regardless of whether it is in the form of As(III) or As(V). Any remaining As could be attributed especially to coprecipitation with Fe oxyhydroxides, which is supported by microprobe data, or to organic matter in the organic-rich topsoil horizons. Based on the XRD results, poorly crystalline goethite and less hematite are the prevalent Fe oxyhydroxide and oxide, respectively, in the soils and are also well known as strong adsorbents for As(V) (Manning and Goldberg, 1996; Gao and Mucci, 2001; Dixit and Hering, 2003; Guo et al., 2007). Poorly crystalline goethite always predominates over hematite in the PS1, PS21 and PS3 soil profiles and hematite was not even detected in the soil underlying the waste dump (PD2 profile). These findings are in accordance with the experimental observations, indicating that preferential formation of hematite occurs at near neutral pH values (7–8) and decreases with increasing or decreasing pH, reaching a minimum in acidic (pH 4) conditions similar to the pH in the PD2 profile (Schwertmann and Murad, 1983; Cudennec and Lecerf, 2006). In addition to crystalline and poorly crystalline Fe oxyhydroxides and oxides, the results of oxalate extraction and PHREEQC modelling predict the presence of amorphous Fe oxyhydroxide, ferrihydrite. The pronounced occurrence of As-bearing ferrihydrite is supported by the facts that goethite or hematite should be formed from the ferrihydrite precursors and this transformation is relatively slow (Schwertmann and Murad, 1983; Cudennec and Lecerf, 2006; Das et al., 2011) and could be even retarded by the presence of As(V) (Paige et al., 1996; Sun et al., 1999; Ford, 2002; Majzlan, 2011). Based on the above hypotheses, the rate of ferrihydrite transformation should be the slowest in the upper soil horizons, in which the As/Fe ratio in oxalate extractions corresponds to the As-richest Fe oxyhydroxides (2.3–13.9 g As kg − 1 active Fe oxyhydroxides). Indeed, these horizons yielded increased amounts of adsorbed As (determined by considering the sorbent as the sum of Fe liberated by the oxalate extraction and adsorbed As, as determined from phosphate extraction), which could be attributed

553

to an increased number of sorption sites provided by amorphous Fe oxyhydroxides. However, the XRD patterns of Fe oxyhydroxidesrich heavy mineral fractions and individual loose Fe oxyhydroxides grains from these fractions did not show any broader diffusive peaks typical for amorphous Fe oxyhydroxides. Thus, the amount of ferrihydrite in the soil probably does not correspond to all oxalate-extractable Fe and some fraction of As and Fe in the oxalate extraction is attributed to alteration of poorly crystalline goethite (Schwertmann, 1973; McCarty et al., 1998). Although the transformation of the ferrihydrite precursor into goethite and hematite results in a decrease in the surface area and an associated reduction in the ability to adsorb As and other contaminants, the dissolved concentrations of arsenic are very low and do not exceed the level of 15 μg L − 1 As (V), which was the highest observed dissolved concentration in the soil pore water (Table 3). This confirms that, during phase transformation, As(V) remains adsorbed or is rather incorporated into the structure of the crystalline product (Pedersen et al., 2006). Arsenate speciation is pH-dependent with H3AsO4, H2AsO4 −, HAsO4 2−, and AsO4 3− having pKa 1 = 2.30, pKa 2 = 6.99, and pKa 3 = 11.88 (Nordstrom and Archer, 2003), respectively. Protonated surface functional groups on Fe oxyhydroxides also undergo pHdependent dissociation. Consequently, pore water-soil partitioning of As controlled by adsorption and desorption are pH-dependent. Arsenate adsorption maxima on Fe oxyhydroxides occurred over a wide pH range, but were highest at pH values below 8 (Stollenwerk, 2003 and references therein). Laboratory investigations of As adsorption on Fe oxyhydroxides in a lower pH range (3–6) showed that As(V) sorption on goethite decreases slightly with increasing pH (Pierce and Moore, 1982; Dixit and Hering, 2003; Giménez et al., 2007). This behaviour probably plays a crucial role in the partitioning of As between the soil and the pore water at the study site. The role of other potential adsorbents, such as clay minerals, Al oxyhydroxides and organic matter, was not apparent and is probably negligible in the multi-component adsorbent system of the studied soil with respect to active Fe oxyhydroxides. 4.2. Long-term perspective on As mobility The results of PHREEQC geochemical modelling support the conclusion that poorly crystalline goethite and ferrihydrite as a major active Fe oxyhydroxides mineral in the Giftkies soil could play an important role in As(V) adsorption. The amount of active Fe oxyhydroxides (based on oxalate extraction) in the upper soil horizons (0.4 m) underlying the waste dump has been estimated at a level of approx. 60000 g m − 3. According to the geochemical modelling, only ~3% of this amount has been equilibrated with As and other contaminants released from the waste seepage flux over the past 350 years. Consequently, the time necessary to equilibrate all the active Fe oxyhydroxides in the soil could be estimated at ~11600 years, assuming that the reactive soil cover is 0.4 m thick and that the annual input of As and other contaminants from the mine waste dump into the soil is similar to that used in the modelling. Our results predict very low mobility of As in the soil at the Giftkies site over long periods of time. It has been shown that As partitioning between the pore water and the soil is pH-sensitive and has no relation to the total As in the soil. These conclusions are in accordance with the results of Erbanová et al. (2008), who proposed a link between higher aqueous As in runoff in three forested catchments in Northern Bohemia, with recovery from acidification. The acidity of atmospheric deposition at the site, caused mainly by SOx and NOx emissions from power plants, increased in 1950–1987, and has been decreasing ever since (Fottová, 2003). In recent years, the pH of spruce throughfall at the nearby Jezeří catchment is circumneutral (Erbanová et al., 2008), whereas the pH of the soil is still substantially lower (Fig. 2). The delay in the high pH signals from atmospheric input to soil-water system could be caused by soil

554

P. Drahota et al. / Science of the Total Environment 414 (2012) 546–555

buffering. Depending on the buffering capacity of soil, and thus its ability to increase the pH, and on the rate of the buffering reactions, we can expect an increase in the soil pH that could cause increased solubility of As in the soil. However, a slight increase in soluble As and its potential migration will probably not represent an environmental problem at the study site since it did not cause either spatial or vertical migration within the soil profile. 5. Conclusions This study shows the importance of active Fe oxyhydroxides in longterm sequestration of As in the contaminated soils near a historical mining waste dump. In spite of the probably diverse nature of the As contamination at this site (smelting of As-rich ores and seepage flux from historical As-rich waste dump), As is ultimately associated with Fe oxyhydroxides (up to 3.2 wt.% As) in the upper soil mineralized horizons as adsorbed and/or coprecipitated species. The efficiency of these processes to limit the As migration, as expressed by low As(V) (b15 μg L − 1) in the soil pore water, cannot be proposed for the other contaminants Cu, Pb and Zn, provided the soil condition is slightly acidic and oxidizing. Based on the XRD measurements, As-bearing compounds are all poorly crystalline goethites with fewer hematites. At slightly acidic pH of the soil, the modelling results indicate that the trapping of the hazardous elements from the percolating waste dump solution by adsorption onto poorly crystalline goethite and amorphous Fe oxyhydroxides is highly effective for As(V) but less effective for metal cations (Cu, Pb and Zn). Quantification of the attenuation capacity of Fe oxyhydroxides in the soil indicates that only ~3% of the active Fe oxyhydroxides has been equilibrated with percolating solutions from the waste dump. As a result, a long-term (~11600 years) natural attenuation of released As could be predicted. Supplementary materials related to this article can be found online at doi:10.1016/j.scitotenv.2011.11.003. Acknowledgements This work was carried out with financial assistance from project KJB300130702 of the Grant Agency of the Academy of Sciences of the Czech Republic. The institutional funding was provided by the Ministry of Education, Youth and Sports MSM 0021620855 and by the Institution Research Plan No. AVOZ30130516. Dr. Vlasta Böhmová and Dr. Martin Racek assisted with microprobe and SEM/EDS measurements; Lenka Jílková is thanked for prompt anion analyses. Dr. Madeleine Štulíková is thanked for correction of the English in the paper. The authors really appreciate all the detailed review of two anonymous referees for improving the manuscript. References Bluteau MC, Demopoulos GP. The incongruent dissolution of scorodite — solubility kinetics and mechanism. Hydrometallurgy 2007;87:163–77. Bufka A, Velebil D. Historical silver ore field district Boží dar (Gottesgab) near Jáchymov (Joachimsthal). Krušné Hory Mts. (Erzgebirge Mts.), Czech Republic. Bull Mineral-Petrol Odd Nár Muz 2005;13:46–61. (in Czech). Cancès B, Juillot F, Laperche V, Polya D, Vaughan DJ, Hazemann J-L, et al. Changes in arsenic speciation through a contaminated soil profile: a XAS based study. Sci Total Environ 2008;397:178–89. Cornell RM, Schwertmann U. The iron oxides. Wiley VCH; 1996. Costagliola P, Benvenuti M, Chiarantini L, Bianchi S, Di Benedetto F, Paolieri M, et al. Impact of ancient metal smelting on arsenic pollution in the Pecora River Valley, Southern Tuscany, Italy. Appl Geochem 2008;23:1241–59. Cudennec Y, Lecerf A. The transformation of ferrihydrite into goethite and hematite, revisited. J Solid State Chem 2006;179:716–22. Das S, Hendry MJ, Essilfie-Dughan J. Transformation of two-line ferrihydrite to goethite and hematite as a function of pH and temperature. Environ Sci Technol 2011;45: 268–75. Dixit S, Hering JG. Comparison of arsenic(V) and arsenic(III) sorption onto iron oxide minerals: implications for arsenic mobility. Environ Sci Technol 2003;37:4182–9. Drahota P, Filippi M. Secondary arsenic minerals in the environment: a review. Environ Int 2009;35:1243–55.

Drahota P, Rohovec J, Filippi M, Mihaljevič M, Rychlovský P, Červený V, et al. Mineralogical and geochemical controls of arsenic speciation and mobility under different redox conditions in soil, sediment and water at the Mokrsko-West gold deposit, Czech Republic. Sci Total Environ 2009;407:3372–84. Dzombak DA, Morel FMM. Surface complexation modelling-hydrous ferric oxide. New York: Wiley; 1990. Erbanová L, Novák M, Fottová D, Doušová B. Export of arsenic from forested catchments under easing atmospheric pollution. Environ Sci Technol 2008;42:7187–92. Filippi M, Doušová B, Machovič V. Mineralogical speciation of arsenic in soils above the Mokrsko-west gold deposit, Czech Republic. Geoderma 2007;139:154–70. Filippi M, Machovič V, Drahota P, Böhmová V. Raman microspectroscopy as a valuable additional method to X-ray diffraction and electron microscope/microprobe analysis in study of iron arsenates in environmental samples. Appl Spectrosc 2009;63:621–6. Filippi M, Drahota P, Machovič V, Böhmová V, Mihaljevič M. Arsenic mineralogy and mobility in the arsenic-rich historical mine waste dump. in preparation. Ford RG. Rates of hydrous ferric oxide crystallization and the influence on coprecipitated arsenate. Environ Sci Technol 2002;36:2459–63. Fottová D. Trends in sulphur and nitrogen deposition fluxes in the GEOMON network, Czech Republic, between 1994 and 2000. Water Air Soil Pollut 2003;150:73–87. Gao Y, Mucci A. Acid base reactions, phosphate and arsenate complexation, and their competitive adsorption at the surface of goethite in 0.7 M NaCl solution. Geochim Cosmochim Acta 2001;65:2361–78. Giménez J, Martínez M, De Pablo J, Rovira M, Duro L. Arsenic sorption onto natural hematite, magnetite, and goethite. J Hazard Mater 2007;141:575–80. Guo H, Stűben D, Berner Z. Removal of arsenic from aqueous solution by natural siderite and hematite. Appl Geochem 2007;22:1039–51. Hruška J, Krám P. Modelling long-term changes in stream water and soil chemistry in catchments with contrasting vulnerability to acidification (Lysina and Luhuv Bor, Czech Republic). Hydrol Earth Syst Sci 2003;7:529–39. ICDD. JCPDS PDF-2 database. Newton Square, PA, USA: ICDD; 2002. Jackson ML, Lim CH, Zelazny LW. Oxides, hydroxides and aluminosilicates. In: Klute A, editor. Methods of soil analysis, part I, physical and mineralogical methodsAmer Soc Agron Inc; 1986. p. 101–50. Keon NE, Swartz CH, Brabander DJ, Harvey C, Hemond HF. Validation of an arsenic sequential method for evaluating mobility in sediments. Environ Sci Technol 2001;35:2778–84. Langmuir D, Mahoney J, Rowson J. Solubility products of amorphous ferric arsenate and crystalline scorodite FeAsO4·2H2O and their application to arsenic behaviour in buried mine tailings. Geochim Cosmochim Acta 2006;70:2942–56. Majzlan J. Thermodynamic stabilization of hydrous ferric oxide by adsorption of phosphate and arsenate. Environ Sci Technol 2011;45:4726–32. Manning BA, Goldberg S. Modeling competitive adsorption of arsenate with phosphate and molybdate on oxide minerals. Soil Sci Soc Am J 1996;60:121–31. Matschullat J. Arsenic in the geosphere — a review. Sci Total Environ 2000;249:297–312. McCarty DK, Moore JN, Marcus WA. Mineralogy and trace element association in an acid mine drainage iron oxide precipitate; comparison of selective extractions. Appl Geochem 1998;13:165–76. Morin G, Calas G. Arsenic in soils, mine tailings, and former industrial sites. Elements 2006;2:97-101. Nordstrom DK, Archer DG. Arsenic thermodynamic data and environmental geochemistry. In: Welch AH, Stollenwerk KG, editors. Arsenic in ground water. Boston: Kluwer Academic Publisher; 2003. p. 1-26. Ondruš P, Veselovský F, Gabašová A, Hloušek J, Šrein V. Geology and hydrothermal vein system of the Jáchymov (Joachimsthal) ore district. J Czech Geol Soc 2003;48:3-18. Oulehle F, Hofmeister J, Hruška J. Modeling of the long-term effect of tree species (Norway spruce and European beech) on soil acidification in the Ore Mountains. Ecol Model 2007;204:359–71. Paige CR, Snodgrass WJ, Nicholson RV, Sharer JM. The crystallization of arsenatecontaminated iron hydroxide solids at high pH. Wat Environ Res 1996;68: 981–7. Paktunc D, Bruggeman K. Solubility of nanocrystalline scorodite and amorphous ferric arsenate: implications for stabilization of arsenic in mine wastes. Appl Geochem 2010;25:674–83. Pansu M, Gautheyrou J. Handbook of soil analysis-mineralogical, organic and inorganic methods. Springer; 2006. Parfitt RL. Optimum conditions for extraction of Al, Fe and Si from soils with acid oxalate. Commun Soil Sci Plant Anal 1989;20:801–16. Parkhurst DL, Appelo CAJ. User's guide to PHREEQC (version 2)— a computer program for speciation batch-reaction, one dimensional transport and inverse geochemical calculations. Denver, Colorado. U.S. Geological Survey report; 1999. p. 99-4259. Pedersen HD, Postma D, Jakobsen R. Release of arsenic associated with the reduction and transformation of iron oxides. Geochim Cosmochim Acta 2006;70:4116–29. Pierce ML, Moore CB. Adsorption of arsenite and arsenate on amorphous iron hydroxide. Water Res 1982;16:1247–53. Richter P, Sguel R, Ahumada I, Verdugo R, Narváez J, Shibata Y. Arsenic speciation in environmental samples of a mining impacted sector of central Chile. J Chil Chem Soc 2004;49:333–40. Schwertmann U. Use of oxalate for Fe extraction from soils. Can J Soil Sci 1973;53: 244–6. Schwertmann U, Murad E. Effect of pH on the formation of goethite and hematite from ferrihydrite. Clay Clay Miner 1983;31:277–84. Shotyk W, Cherburkin AK, Appleby PG, Fankhauser A, Kramers JD. Two thousand years of atmospheric arsenic, antimony, and lead deposition recorded in an ombrotrophic peat bog profile, Jura Mountains, Switzerland. Earth Planet Sci Lett 1996;145:E1–7.

P. Drahota et al. / Science of the Total Environment 414 (2012) 546–555 Smedley PR, Kinniburgh DG. A review of the source, behaviour and distribution of arsenic in natural waters. Appl Geochem 2002;17:517–68. Smith E, Naidu R, Alston AM. Arsenic in the soil environment: a review. Adv Agron 1998;64:149–95. Stollenwerk KG. Geochemical processes controlling transport of arsenic in groundwater: a review of adsorption. In: Welch AH, Stollenwerk KG, editors. Arsenic in Ground Water: Geochemistry and Occurrence. Boston: Kluwer Academic Publisher; 2003. p. 67-100. Sun X, Paige CR, Snodgrass WJ. Combined effect of arsenic and cadmium on the transformation of ferrihydrite onto crystalline products. J Univ Sci Technol B 1999;3: 168–73. Tokoro C, Yatsugi Y, Koga H, Owada S. Sorption mechanisms of arsenate during coprecipitation with ferrihydrite in aqueous solution. Environ Sci Technol 2010;44: 638–43. UNIQUANT 4. Software for standardless X-ray fluorescence analysis. The Netherlands: Omega Data System bv; 1999.

555

Van Herreweghe S, Swennen R, Vandecasteele C, Cappuyns V. Solid phase speciation of arsenic by sequential extraction in standard reference materials and industrially contaminated soil samples. Environ Pollut 2003;122:232–42. Wenzel WW, Sletten RS, Brandstetter A, Wieshammer G, Stingeder G. Adsorption of trace metals by tension lysimeters: nylon membrane vs. porous ceramic cup. J Environ Qual 1997;26:1430–4. Wenzel WW, Kirchbaumer N, Prohaska T, Stingeder G, Lombi E, Adriano DC. Arsenic fractination in soils using an improved sequential extraction procedure. Anal Chim Acta 2001;436:309–23. West S, Charman DJ, Grattan JP, Cherbukin AK. Heavy metals in Holocene peats from south west England: detecting mining impacts and atmospheric pollution. Water Air Soil Pollut 1997;3–4:343–53.