Mercury mobility in mine waste from Hg-mining areas in Almería, Andalusia (Se Spain)

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Mercury mobility in mine waste from Hg-mining areas in Almería,Andalusia (Se Spain)☆

Andrés Navarro a,⁎, Esteve Cardellach b, Mercé Corbella b

a Dep. Mec. de Fluidos. Universitat Politècnica de Catalunya, ETSEIAT, Colón 7-11, 08222-Terrassa (Spain)b Dep. Geologia, Universitat Autònoma de Barcelona, 08193 Bellaterra, Spain

a b s t r a c ta r t i c l e i n f o

Article history:Received 10 April 2008Accepted 11 August 2008Available online 11 September 2008

Keywords:MercuryMine wastesSpeciationContaminationMINTEQ

Mining wastes and calcines from two abandoned mining areas (Valle del Azogue and Bayarque in Almería)have been characterized. In the mining wastes, the dominant mercury phases are cinnabar and elementalmercury in the matrix. In the calcines, however, the dominant mercury phase is elemental mercury boundedto the matrix. Water-leaching experiments were conducted on low-grade stockpiles and calcines in order tosimulate the mobilization of mercury by runoff under environmental conditions. The laboratory column-leaching experiments show a possible mobilization of mercury from Hg0 dissolution, colloid transport and apossible dissolution of calomel and other soluble phases in the mine wastes from the Valle del Azogue andBayarque mines. Equilibrium speciation modeling of Hg, conducted using the numerical code MINTEQ,showed that the theoretical dominant mercury species in the calcine and mining wastes samples are Hg(OH)2, HgCl2, HgClOH and Hg0. In some leachates obtained from the Valle del Azogue mining wastes (sampleA06), the high Hg concentrations may indicate the possible dissolution of mineral phases such as calomeland other soluble phases, which are subsaturated. The environmental results indicate a great environmentalmobility of mercury, especially during wet episodes associated with intense precipitation events, when thereare significative amounts of secondary soluble minerals.

© 2008 Elsevier B.V. All rights reserved.

1. Introduction

The release of mercury into the environment in Hg-mining areas isgenerally associated with the abandonment of mine wastes, which ismainly composed of calcines (waste originated in themetallurgy of Hg).Mine waste impoundments that contain waste rock and low-gradestockpiles (Rytuba, 2003; Higueras et al., 2003; Wang et al., 2004;Fernández-Martínez et al., 2005; Qiu et al., 2005; Higueras et al., 2006;Navarro et al., 2006) and mine waste from gold-mining activities(Churchill et al., 2004; Shaw et al., 2006) may produce themobilizationof mercury to the soil, groundwater, surface water and streamsediments. The most abundant Hg deposits are the silica-carbonateand hot-spring deposits, where cinnabar (HgS) is the main ore mineral.In SE Spain, themainHgmineral deposits are located in the Betic Rangesand consist of low-sulfidation epithermal hot-spring deposits such asthe Valle del Azogue mine (Viladevall et al., 1999; Navarro et al., 2000,2006; Mendoza et al., 2006) and stratabound deposits such as theBayarque mine.

Mercury mine waste and Hg-enriched soils are a potential sourceof particulate and soluble Hg species (Rytuba, 2005), which may betransported from contaminated areas as Hg0 vapor in semi-aridenvironments (Navarro et al., 2000; Gustin et al., 2002), as ionicsoluble phases or as colloid particles (Shaw et al., 2001; Lowry et al.,2004). Old mining sites treated ore by roasting it in furnaces withcondensing systems. This produced calcines and secondary Hg and Feminerals (mainly metacinnabar and Fe oxides), which were dumpednear the metallurgical facilities. The concentration of Hg in calcinesvaries greatly, but is usually greater than 1000 ppm (Gosar et al., 1997;Rytuba, 2000; Rytuba et al., 2001; Gray et al., 2002; Gray, 2003; Loredoet al., 2003; Zhang et al., 2004). In this study, the concentration of Hgin calcines ranged from 330 to 1000 ppm.

The mobilization of Hg surrounding mining areas in semi-aridenvironments such as the study area, is mainly caused by atmosphericemissions, by mechanical dispersion of particles (wind transport) andby water-leaching from the mine waste deposits. In aquatic environ-ments, hydroxide, chloride and sulfide may control speciation, and themobilization of mercury may be influenced by processes related to theaqueous phase (precipitation and dissolution of solids, complexformation and redox reactions) and processes related to interactionwith associated solidmedia (sorption reactions and colloidal transport).

Cinnabar is very insoluble and much less volatile and leachablefrom soils than other forms of Hg, but it can dissolve in the presence oforganicmatter bymeans of surface complexation (Ravichandran et al.,

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☆ In this paper mine wastes are classified as mining wastes (overburden, waste rockand low-grade stockpiles), processing wastes (tailings, sludges and waste water) andmetallurgical wastes (calcines and slags).⁎ Corresponding author. Tel.: +34 93 7398151; fax: +34 93 7398101.

E-mail address: [email protected] (A. Navarro).

0375-6742/$ – see front matter © 2008 Elsevier B.V. All rights reserved.doi:10.1016/j.gexplo.2008.08.004

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1998, 2003). In surface environments, such as surface soils orgroundwater, metallic Hg0 may be an stable species (Davis et al.,1997; Morel et al., 1998; Lechler, 1999).

In oxidizing environments, the most soluble inorganic species areHgCl2 at low pH, HgClOH at neutral pH, and Hg(OH)2 at high pH.Mercurous compounds (Hg+), which rapidly oxidize to mercuric forms(Hg2+), and mercuric cations may be absorbed by clay minerals, oxidesand organic matter, depending on the pH. Metallic Hg0 can be sorbed tomineral matrix components (Biester et al., 1999; Navarro et al., 2006)suchasdolomite, ironoxyhydroxides andhumic acids derived fromsoilsassociated with contaminated areas. Besides, organic matter is animportant factor in controlling mercury sorption in soils and sedimentsdominated by Hg-hydroxy species (Ying et al., 1996).

Colloidal transport of mercury from mine waste has been demon-strated in column experiments, where the leaching of calcines indicatesthe transport of Hg-bearing colloids of 0.05–0.4 μm of alunite–jarosite,hematite and amorphous silica–aluminum (Shaw et al., 2001; Rytuba,2002, 2003; Lowry et al., 2004), reaching Hg concentrations of 2 mg/L.

Also, Gray et al. (2000) reported high levels of mercury in calcines(2000 ppm), stream sediments (170 ppm) and experimental leachates(1500 μg/L) associatedwith the presence of soluble phases ofmercuryor colloidal material.

The main objectives of this study were: (1) to characterize thegeochemical composition and the Hg species in the mine wastes of theValle del Azogue and Bayarque mines, (2) to evaluate the aqueousenvironmental mobility of mercury from leaching experiments, using

three characteristics mine wastes samples: a sample of calcine, wherethe dominant species is Hg0 (M03), a sample of mining wastes (A06)where may exist Hg soluble phases (calomel and kuzminite), and asample of mining wastes where the mercury dominant species iscinnabar (B03), and (3) to study the influence of the speciation in themobilization of Hg in a semi-arid environment through geochemicalmodeling of the leaching experiments.

2. Study area

The mine waste samples used in the column experiments wereobtained from the oldmines in Valle del Azogue and Bayarque, locatednorth of Almería in the SE of the Betic Ranges (Fig. 1). In the 19thcentury, the Valle del Azogue mine was the main mercury mine in theBetic Ranges. It was active from approximately 1873 to 1888, bymeansof underground works and small open pits located near two smeltersites (Becker, 1888), producing unknown ore quantities (Fig. 2). Themineralization consists of a set of small veins and breccias cuttingacross the metamorphic host rocks from the Permian-Triassic period(phyllites and marbles of the Alpujárride Complex) and occasionallyoverlying marls, sandy marls and limestones from the Tertiary period.The mineral paragenesis comprises stibnite, cinnabar, As minerals(realgar and orpiment), sphalerite, siderite, chalcopyrite, pyrite,quartz, calcite and baryte (Martinez et al., 1997; Navarro et al.,2006). Hydrothermal and supergene alteration of primary mineralsresulted in an association of secondary Fe–Mn and Sb–As oxides and

Fig. 1. Location map (a) and synthetic geology of the study area. NQ: Neogene-Quaternary, UM: volcanic–shoshonitic rocks, PT: metamorphic basement, F: main faults.

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hydroxides, kaolinite, jarosite and gypsum. Other mineral phasesdetected (Mendoza et al., 2006) include metacinnabar (HgS), Se–HgS,tiemannite (HgSe), corderoite (Hg3S2Cl2), shakhovite (Hg4SbO5(OH)3)and schuetteite (Hg3(SO4)O2). In a preliminary study, the distributionof mercury in soils and the measurements of Hg in soil gas indicatedsignificant anomalies up to 1 km in length and 120 m in width,associated with high contents of As, Sb and Pb (Viladevall et al., 1999).Also, the mercury content in vegetation ranges between 5 ppm and54 ppm in thyme (Thymus vulgaris), and 70 ppm in vegetables. The

numerical modeling of the mercury vapor transport showed that themobilization of Hg seems conditioned by the volatilization of Hg0,which was evaluated by means of a single-component advection-dispersion model (Navarro et al., 2000). The applied numerical modelsuggested that gas fluxes in the soil can be produced by a dominantdiffusion mechanism. This is a consequence of the presence ofelemental Hg coming from the vapor emitted by the chimney of theroaster facility, and the presence of elemental Hg bounded to the soilmatrix. The research of solid-phase speciation of Hg from the Valle del

Table 1Metal and metalloid concentrations of mine waste and soils in the Valle del Azogue mine.

Description Au Ag As Ba Cd Cr Cu Hg Fe Pb Sb Sr V Zn

SampleAZ21 Calcine 10 10 448 3.6 5.8 110 48 330 2.49 549 1350 843 82 2932AZ22 MW 13 b0.3 12.9 0.09 b0.3 90 18 b1 4.37 44 20.8 285 128 117AZ23 MW b2 b0.3 16.7 0.07 b0.3 50 52 6 2.23 30 8 1134 46 70AZ24 Calcine 150 76.6 1650 11.0 9.7 161 32 1010 4.18 1873 22200 807 31 3528M04 MW 40 50 864 11.0 2 b30 8 4000 1.28 549 7460 741 22 465M05 MW 5 0.9 129 0.25 b0.3 115 10 50 5.13 45 200 231 127 648M06 MW 5 8.7 331 4.5 29.9 83 8 399 1.27 303 1880 614 75 8524M07 MW 63 0.7 233 0.16 b0.3 108 208 23 4.66 28 66 102 111 97A01 MW 23 9.6 300 7.5 0.9 100 16 530 1.59 489 1200 1157 73 336A02 MW 32 24.6 620 5.8 13.9 63 35 1000 2.37 1497 1800 727 59 4633A03 MW+soil 7 3.1 184 2.9 1.9 65 31 210 3.69 323 625 1078 83 543A04 Calcine 164 108.7 680 17.0 18.4 b50 143 540 3.5 3422 8600 546 20 14295A05 Calcine 66 39.3 320 10.0 3.9 120 32 400 2.39 1143 4200 896 22 878A06 MW 35 9.7 610 1.7 2.1 97 43 600 2.96 778 2200 1013 79 2074A07 MW 43 21.3 410 3.8 0.7 96 31 820 3.32 799 2600 1630 59 954M03 Calcine 76 75.1 1610 17.0 4.3 b50 46 470 3.78 2555 14800 3321 34 2002

WastesMean 45.8 27.3 526.1 6.0 5.7 70.5 47.5 649.1 3.0 901.6 4325.6 945.3 65.6 2631Min b2 b0.3 12.9 0.07 b0.3 b50 8.0 b1 1.27 28.0 8.0 102.0 20.0 70.0Max 164 108.7 1650 17.0 29.9 161 208.0 4000 5.13 3422 22200 3321 128 14295

SoilMean 43.2 4.5 299.4 0.89 6.5 99.7 23.1 348.0 3.4 222.4 2669.3 446.8 – 434Min 5.0 b0.2 18.0 0.02 0.5 10.0 4.0 1.0 0.6 7.0 19.0 0.5 90 50.0Max 480.0 30.8 1200 5.3 106 430 55.0 2300 6.48 1308 32000 1500 90. 2970DMMAS-101 618 2040 0.0732 161 7.64 14.3DMMAS-101 cert 612 2087 0.0475 144 7.44 14.7GXR-4 3.4 b0.3 6488 0.111 51 221 89 80GXR-4 cert 4 0.86 6520 0.110 52 221 87 73NCD 11.0 b0.3 24.1 0.08 0.8 35 12 b1 1.08 19 2.1 300 43 4NIL (⁎) – 15⁎⁎ 55 0.0625 12 380 190 10 – 530 15 – 250⁎⁎ 720

Values in mg kg−1 except Ba and Fe (%) and Au (μg kg−1). MW: mining wastes, NIL (⁎): The Netherlands soil intervention values, (⁎⁎): indicative level of contamination. NCD: non-contaminated soils.

Fig. 2. The Valle del Azogue mine. (a) Calcines and mining wastes, (b) condensation system of old furnaces, (c) soil and mine wastes sampling points.

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Azogue soils by SEM and solid-phase-Hg-thermo-desorption (SPTD)indicated that the predominant Hg species was cinnabar and metallicHg0 in thematrix (Navarro et al., 2006). Besides, the aqueousmobility ofHg in the contaminated soils may be originated from the dissolution ofHg0 present in the soil, reaching concentrations close to solubility ofelemental Hg at environmental conditions (Navarro et al., 2006).

The Bayarquemine is located near the valley of the Almanzora Riverand a town of the same name. Like the Valle del Azogue mine, thismineralization is locatedwithin theAlpujárride Complex in the SE of theBetic Ranges, hosted in the Higher and Middle units (Booth-Rea et al.,2003). The stratigraphic succession of the mineralized area includes aphyllitic formation made up of green and purple phyllites, quartzites,limestones and metabasites, below the level of the massive limestonesthat make up the upper level of the Alpujárride Complex. Themineralization mainly consists of quartzites and carbonatic rocks,which both occur as disseminations, and quartz-phyllitic clots andpockets of cinnabar. The mineralogy mainly consists of cinnabar,malachite, azurite, hematites, galena, goethite, siderite, chalcopyrite,bornite, chalcosine, covellite andfluorite. The deposit is similar to that ofNew Idria mine in California, although it also resembles the meta-morphic-hostedmercurydeposits of Levigliani in Italy (Dini et al., 2001).Nevertheless, not much is known about this type of deposit. Theunderground ore of Bayarque was mined by means of drift mining andtreated in roasting furnaces, where the cinnabar was oxidized and SO2

andHg0 vaporwasproduced. The Bayarquemine operated,mainly, from1966 to 1973, producing approximately 33000 t of ore (0.6–42.0% Hg),and there has been no mining activity since 1973. The mine waste andthe old industrial facilities of the Valle del Azogue and Bayarque mineshave remained in the same place and no reclamation work has beendone (Fig. 2).

3. Materials and methods

In order to evaluate the mobility of Hg from mine wastes and tocomplete the geochemistry of the Valle del Azogue area, superficialmine wastes sampling was carried out closely to the roaster facilityand the main mine wastes impoundments, between 2004 and 2007(Fig. 2c). Samples M03 and A01 to A07 were obtained in 2004 and theremainder samples were obtained between 2006 and 2007. Thelocation of samples (Fig. 2c) was conditioned by the situation ofcalcine and mining wastes impoundments. Thus, samples M03 andA01 to A07 were located along the mining area, comprising calcines(M03, A04 and A05) and mining wastes (A01 t0 A03, A06 and A07).Furthermore, samplesM04 to M07 (mining wastes) were located nearthe main open pit in the Southern area, and AZ21 to AZ24 (calcines:AZ21 and AZ24) near the main calcine dumping area (Fig. 2b and c).The duration of roasting activities in the Valle del Azogue mine was,approximately, estimated in 15 years, concluding about 1888 (Becker,1888). In the Bayarque mine were sampled calcines, mining wastesand an alluvial soil, downstream the mining area and near the mainmining wastes impoundments.

The sampling method used was derived from standard methodsfor describing and sampling contaminated soils (EPA, 1991). Thesamples were sealed to minimize exposure to atmospheric gasses anddried at 30 °C for 48 h. Before the chemical analyses, the calcines and

mine waste were ground in a mechanical mortar and homogenized.Au, Ag, As, Ba, Br, Cr, Fe, Hg and Sb, were quantitatively analyzed byInstrumental Neutron Activation Analysis (INAA) and Cd, Cu, Pb, Zn,Ag, Sr, and V were quantitatively analyzed by Inductively CoupledPlasma Emission Spectroscopy (ICP-OES) in Actlabs (Ontario, Canada).

Quality control for the Hg andmetal determinations in solid sampleswas addressed with method blanks, duplicates and certified referencematerials: DMMAS-101 for INAA andAN-G, SDC-1, DNC-1, SCO-1, GXR-1,GXR-2, GXR-4 andGXR-6 for total digestion and ICP-OES analysis. Limitsof determinationwere 1 ppm for total Hg, although some samples wereanalyzed by cold vapor-FIMS with a limit of 0.5 ppb. The analysis ofstandard reference materials showed small differences between thesamples and certified materials (Tables 1 and 2).

The mine waste samples were studied using transmitted andreflected light microscopy, X-ray diffraction (XRD) and scanningelectron microscopy (SEM) with an attached energy-dispersive X-rayspectroscopy (EDS) system. These techniques allowed us to identifythe mineral phases and analyze the major and trace element contentsof the most significant minerals. SEM and EDS were used tocharacterize the solid forms of Hg in ore and calcine samples. Theanalyses were performed at the Electron Microscopy Laboratory of theAutonomous University of Barcelona.

The Hg phases were determined by solid-phase-Hg-thermo-desorption (SPTD), which is based on the specific thermal desorptionor decomposition of Hg compounds from solids at different tempera-tures (Biester and Schulz, 1997; Navarro et al., 2006). Hg thermo-desorption curves were determined using an in-house apparatus thatconsisted of an electronically controlled heating unit and an Hgdetection unit. For Hg detection, a quartz cuvette, through which thethermally released Hg is purged, is placed in the optical system of anatomic absorption spectrometer (Perkin–Elmer AAS 3030). The resultsare depicted as Hg thermo-desorption curves showing the release ofHg over temperature.

Five leaching tests were performed to evaluate metal mobility fromminingwastes and calcines, using a columnwith an internal diameter of150 mm, a length of 751 mm and an endpiece with a 0.50 μm filter. Thecolumn was uniformly packed with waste material between two thin

Table 3Majoritary identified minerals in the samples used in the lixiviation tests.

Sample Identified minerals

M03 QuartzBariteIllite

A06 QuartzBariteDolomiteCalciteHuntiteIlliteInyoite

B03 QuartzDolomiteCalciteChlorite–serpentineIllite

Table 2Geochemical composition of samples obtained in the Bayarque mine.

Description Au Ag As Ba Cd Cr Cu Hg Fe Mn Pb Sb Sr V Zn

SamplesB01 Calcine 9 b0.3 9.6 380 1 77 24 66 3.0 485 20 6 344 83 77B02 Soil 8 b0.3 20.8 460 0.3 91 51 33 3.41 731 33 9.1 233 107 103B03 Mining wastes 5 b0.3 63.6 520 1.3 96 46 400 3.72 743 38 10.1 249 117 87B04 Mining wastes 2 b0.3 5.1 b50 0.3 84 17 4600 3.04 1350 31 34.4 134 15 93

Values in ppm, except Au (ppb).

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covers of very low-reactivity polystyrene particles with an equivalentdiameter of 2.97mm. Low-mineralizationwaterwas introduced into thecolumn by a rain simulator connected to a titration pump, whichprovided a constant flow rate. Leaching experiments were performedusing a stationaryflow rate of 2.4 Lmin−1 for 240min, the equivalent of

approximately 6.7 pore volumes. The resulting Darcy velocity in theexperiment was 3.772·10−5 m/s.

The effluents obtained from the experiments were filtered usingcellulose nitrate films with a diameter of 0.45 μm. After determining thepH, temperature, oxidation–reductionpotential andelectrical conductivity,

Fig. 3. Solid-phase-Hg-thermo-desorption (SPTD) of samples used in the leaching tests. (a) M03 calcine sample. Modified from Navarro et al. (2006). (b) A06 mining wastes sampleof the Valle del Azogue mine. (c) B03 mining wastes sample of Bayarque mine.

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the samples were stored at 4 °C. The pH was measured potentiome-trically and the pH-meter was calibrated before each measurement.Conductivity was determined using a conductimeter calibrated with aNaCl solution. The oxidation–reduction potential was measured usingan ORP meter with a combined platinum electrode and was used asapproximate data in the geochemical modeling.

Sampleswere then acidifiedwith HCl (pH=1.9) and analyzed usinginductively coupled plasma mass spectrometry (ICP-MS) at ACTLABS(Ontario, Canada). Cr, Mn, Fe, Co, Ni, Cu, Zn, As, Se, Ag, Cd, Sb, Ba, Au,Hg, Tl and Pb were determined. Unacidified samples were stored inlow-density polyethylene (LDPE) bottles and Cl, CO3H, CO3, NO3 andSO4 were analyzed using ion chromatography and other commonmethods (APHA, 1989). The accuracy of the results was assessed byexamining blanks and replicates and by performing charge-balancecalculations. Standard reference material NIST 1640 (ICP-MS) wasused to evaluate accuracy.

Geochemical modeling was performed using, first, PHREEQCnumerical code (Parkhurst and Appelo, 1999) and finally MINTEQ(Allison et al., 1991), since their results were more coherent, in aneffort to elucidate the precipitation–dissolution processes involved inthe leachate experiments. It was used to evaluate the hydrogeochem-ical analyses of leachates, evaluate the speciation of dissolvedconstituents of leachates and calculate the saturation state of theeffluents. The MINTEQ thermodynamic database was used for thechemical equilibrium calculations. The mobility of metals was studiedby means of pH–Eh diagrams obtained using the MEDUSA hydro-geochemical code (Puigdomenech, 2004).

4. Results and discussion

4.1. Mine waste geochemistry

Previous data on wastes and soils at the Valle del Azogue mineshow large amounts of As, Ba, Hg, Pb, Sb and Zn, and significant

Fig. 4. SEM image of a calcine particle (sample M03) and EDS spectrumwhich indicatesthe presence of Hg, Ba and Zn.

Fig. 5. SEM image of a mining wastes particle of the Valle del Azogue mine and EDSspectrum. (a) Sample A06 which indicates the presence of calomel. (b) Sample A06which indicates the presence of kuzminite.

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content of Au, Ag, Cd, Cr and V (Viladevall et al., 1999; Navarro et al.,2000). These materials also showed significative amounts of As, Ba,Cd, Cr, Hg, Pb, Sb and Zn that exceeded the maximum allowed by theNetherlands Soil Contamination Guidelines (Table 1). The high Hgconcentrations in calcines may be due to the inefficient and in-complete process of cinnabar retorting, the possible re-adsorption ofHg by calcines into the furnace, and mine waste exposed to stackemissions for several years. The results of the materials used in thecolumn-leaching experiments (M03, A06 and B03 samples), consist-ing of calcines and mining wastes (Tables 1 and 2) indicate highcontents of Ag, As, Ba, Hg, Pb, Sb and Zn, in concordance with previousstudies, although the Bayarque sample (B03) only presents highamounts of Hg, Mn and V.

The very high Hg concentration in calcines and mining wastessuggests that significant amounts of Hg may be released to theenvironment, particularly to soil, fluvial sediments and groundwater,

during the episodicwet periods in this semi-arid area. Indeed, very highconcentrations of Hg (N70 ppm) were also detected in the vegetationsurrounding the Valle del Azogue mining area (Viladevall et al., 1999).

4.2. Mineralogical analysis

The mine waste samples contain a diverse mineralogy, whichconsist of a dominant species and minor phases (Table 3). The mostabundant mineral phases identified by XRD are quartz and barite. Theminor phase minerals (Table 3) are illite, calcite, dolomite, chlorite–serpentine, huntite and inyoite.

The SPTD determinations showed two different temperatureranges where Hg was released from the samples (Fig. 3a, b and c):220–250 °C and 310–330 °C. No “free metallic Hg” was detectedbecause its temperature of detection is about 100 °C (Biester et al.,1997, 2000) and the thermo-desorption curves don't show any peak atthis low temperature (Fig. 3). Based on previous interpretations, thefirst Hg release peak detected indicates Hg released from the matrix,whereas the second peak at higher temperatures indicates theoccurrence of cinnabar (Navarro et al., 2006). This Hg-matrix presencemay be found in calcinated samples because of cinnabar processing,which produces Hg0 that may be re-adsorbed to the calcinatedmaterial during cooling. In mine waste samples A06 and B03(excavated host-rock deposit cover and low-grade ore, respectively),dominant Hg0-matrix and cinnabar were detected.

The detailed SEM and EDS systems study of the mine wastesamples showed the presence of primary and secondary cinnabarassociated with barite, pyrite and botroydal pyrite (Fig. 4). Also, SEMobservations showed several small particles that contain both Hg andCl (Fig. 5a) and that may be associated to calomel (Hg2Cl2). Moreover,some particles that contain both Hg and Br (Fig. 5 b) were observedand that may be associated to kuzminite (Hg2 (Br,Cl)2).

4.3. Leachate geochemistry

Water-leaching experiments were conducted on calcine materials(sample M03) and mining wastes (samples A06 and B03) in order to

Fig. 6. Evolution of Hg concentration eluted from calcine (sample M03) and Bayarquemining-wastes (sample B03) leaching column experiments. Hg-SF: lixiviates withoutfiltering. Hg: filtered lixiviates (0.45 μm). Hg–Bayarque: filtered lixiviates (0.45 μm).

Table 4

Element Time pH Eh Mn Fe Cu As Se Ag Cd Sb Au Hg Pb

Unit Min mV ug/L ug/L ug/L ug/L ug/L ug/L ug/L ug/L ug/L ug/L ug/L

Detection limit 0.1 10 0.2 0.03 0.2 0.2 0.01 0.01 0.002 0.2 0.01

a. Metal composition of unfiltered leachates of sample M03CAL-1 0 8.13 87 197 1970 72.6 N200 2.8 5.5 7.92 N100 0.009 13.2 N200CAL-2 20 8.18 116 63.3 910 24.1 N200 1.6 3.1 3.56 N100 0.007 2.2 N200CAL-3 30 8.22 126 35.4 630 14.5 N200 1.5 3.3 2.07 N100 0.006 4 N200CAL-4 45 8.2 133 26.8 580 11 N200 1.3 4.3 1.66 N100 0.006 6.5 N200CAL-5 60 8.24 137 61.7 1370 23.8 N200 1.5 5.6 3.75 N100 0.004 10 N200CAL-6 90 8.21 134 26 980 10.9 N200 0.9 4.9 1.62 N100 0.006 8.8 N200CAL-7 120 8.23 147 17.5 480 7.8 N200 0.9 4.6 1.17 N100 0.006 5.8 N200CAL-8 180 8.18 134 17.8 430 7.6 N200 0.9 3.2 1.18 N100 0.004 3.5 N200CAL-9 240 8.24 139 8.7 250 7.5 171 0.7 1.8 0.59 N100 0.003 1.2 159CAL-10 312 8.27 137 7.3 260 3.6 149 0.7 2.6 0.5 N100 0.003 2.4 141CAL-11 402 8.27 146 15.4 550 6.4 N200 0.8 6.1 0.98 N100 0.005 9.4 N200

b. Metal composition of unfiltered leachates of sample A06A06-1A 0 7.64 −184 220 2110 53.4 95.2 0.4 67.7 10.2 25.8 0.036 1110 N 200A06-2A 10 7.64 −184 76.6 1110 12.9 59.7 b0.2 37.8 4.03 43.3 0.021 541 95A06-3A 20 7.67 −201 108 1660 17.7 76.6 b0.2 30 4.18 52.4 0.02 502 135A06-4A 31 7.65 −191 62.7 1010 11.6 56.8 b0.2 47.1 5.49 44.7 0.023 913 84.3A06-5A 45 7.56 −139 33.1 70 4.8 10.2 3.9 90.4 15.8 13.5 0.034 N2000 10.2A06-6A 60 7.42 −60 65.7 340 14.4 26.9 23.1 188 38 28.5 0.061 N2000 129A06-7A 90 7.46 −72 43.6 220 8.9 15.4 15.3 141 28.6 17.3 0.035 N2000 60A06-8A 120 7.43 −89 39.2 170 11.1 11.4 7.3 132 25.5 16.4 0.044 N2000 43.8A06-9A 150 7.42 −96 56.5 540 12.6 32 10.5 120 22.3 29.6 0.04 N2000 73.1A06-10A 180 7.37 −60 21.3 30 3.1 6.39 b0.2 77.2 12.8 8.65 0.026 1330 17.9A06-11A 210 7.44 −95 205 1120 39.5 62.6 9.9 112 21.9 19.6 0.044 N2000 160A06-12A 240 7.16 38 19.1 50 3.2 7.84 b 0.2 69 8.78 12.5 0.017 911 10.6

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simulate the mobilization of mercury by runoff in environmentalconditions. In the column experiments, high and variable concentra-tions of mercury were found in the leachates obtained from thecalcine sample (M03) and Bayarque mining wastes sample (B03)(Fig. 6). Besides, large amounts of Hg were found in the unfiltered andfiltered samples from mining wastes leaching of the Valle del Azogue(Tables 4b and 5a).

The leaching experiment on calcine sample M03 showed themobilization of mercury concentrations up to 21.6 μg/L (Fig. 6), valuenear to solubility of Hg0 (Navarro et al., 2006), which is coherent withsolid speciation of this sample, that is mainly composed of Hg0-matrixbound (Fig. 3a).

Fig. 6 shows the minor differences between of Hg concentrationsthe filtered and unfiltered samples of calcine (Valle del Azogue mine)and Bayarque mining wastes. On the contrary, the concentration of Feand Mn in the unfiltered leachates is higher than the respectiveconcentrations in the filtered samples (Fig. 7), especially in the case ofMn. The mobilization of Fe and Mn in the unfiltered samples could beassociated with colloid transport of Fe andMn oxyhydroxide particles.Other contaminants, such as As and Sb, are generally found in larger

amounts, although with lower concentrations than those detected insimilar mine wastes (Gray et al., 2000).

The leachates obtained from sample A06 (Valle del Azogue miningwastes) contained high mercury concentrations, up to 2 mg/L in theunfiltered samples and 1.5 mg/L in the filtered samples (Tables 4band 5; Fig. 8). Large amounts of chloride, sulfate and bicarbonate(Fig. 9) associated with high contents of metals and metalloids werealso detected (Tables 4b and 5a). The presence of large amounts ofchlorides, sulfates and some metals, may indicate the possibledissolution of secondary salts detected in previous studies (Navarroet al., 2006) and the possible presence of mercury soluble phases,since the detected concentrations of this metal were very high. In fact,the solubility of Hg0 at environmental conditions is 56 μg/L, and thesolubility of HgCl2, Hg2Cl2 and Hg2(SO4) are 74, 2 and 600 mg/L,respectively (Davis et al., 1997). Moreover, the SPTD and SEM-EDSdeterminations of sample A06 (Fig. 3b and 5) showed the presence ofHg0-matrix bound, cinnabar and possible soluble phases such ascalomel and kuzminite with large solubilities, which may produce theHg concentrations detected in the leaching tests on A06 miningwastes sample (Tables 4b and 5).

Table 5

Element time pH Eh Mn Fe Cu As Se Ag Cd Sb Au Hg Pb

Unit min mV ug/L ug/L ug/L ug/L ug/L ug/L ug/L ug/L ug/L ug/L ug/L

Detection limit 0.1 10 0.2 0.03 0.2 0.2 0.01 0.01 0.002 0.2 0.01

a. Metal composition of filtered leachates of sample A06AG-1 0 8.14 219 44.2 750 17.4 38 11.1 82.2 3.95 92.5 0.04 716 66.6AG-2 10 8.13 221 136 1010 32.9 55.8 9.6 38.3 4.54 130 0.044 567 134AG-3 20 8.22 218 33.6 520 11.5 38.4 5.4 14.9 1.18 86.5 b 0.02 243 57.9AG-4 30 8.09 224 41.5 890 13.4 50.6 5.1 19 1.68 71.9 0.038 332 76.8AG-5 60 8.11 231 21.6 270 8.5 22.3 5.8 36.3 2.3 28.2 0.024 394 36.4AG-6 90 8.12 222 36.7 810 17.4 39.8 8.9 54.1 4.19 69.2 0.048 685 54.3AG-7 120 8.03 221 47.2 5700 b20 165 26 388 6.95 639 b0.2 1500 353AG-8 180 8.15 212 85.6 1140 25.2 57.6 7.5 28 2.93 147 0.04 543 88

b. Metal composition of filtered leachates of sample B03BA-1 0 8.30 59 429 2640 72.7 27.8 1.6 b0.2 0.76 2.73 0.004 0.8 42.3BA-2 40 8.27 75 86.3 1080 40.8 17.8 1.4 b0.2 0.47 3.77 0.004 0.9 11.8BA-3 90 8.30 89 61.5 880 24.6 12.5 1.1 b0.2 0.24 2.97 0.003 1.3 8.45BA-4 120 8.32 101 30 300 17.6 9.8 1 b0.2 0.18 2.62 0.002 0.7 3.54BA-5 180 8.33 107 23.9 290 15.5 9.12 1 0.2 0.16 2.63 b0.002 1.4 2.83BA-6 240 8.31 109 16.7 200 9.8 5.65 0.9 0.4 0.13 2.33 b0.002 0.6 5.81BA-7 270 8.30 125 14.2 130 6.9 5.16 0.9 0.4 0.12 1.78 b0.002 0.4 1.84

Fig. 7. Evolution of Mn and Fe concentration eluted from calcine leaching columnexperiments on M03 sample. Mn–SF: lixiviates without filtering. Mn: filtered lixiviates(0.45 μm).

Fig. 8. Evolution of Hg concentration eluted from sample A06 (Valle del Azoguemining-wastes) leaching experiments. Hg–SF: lixiviates without filtering. Hg: filtered lixiviates(0.45 μm).

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The concentration of leached mercury in the unfiltered samplesfrom A06 mining wastes (Fig. 8), presents an evolution similar to thatof chloride (Fig. 9). The increase and decrease in the concentrations ofthese two elements coincide throughout the column experiment. Thismay indicate the leaching of soluble mercury associated with Cl−,likely due to the presence of soluble mercury chlorides detected in themineralogical study. The possible presence of these mercury com-pounds and large amounts of mobilized mercury was found by Grayet al. (2002), who detected high concentrations ofmercury (1500 μg/L)in the leachates of calcine samples and attributed it to colloidtransport and water-soluble Hg compounds. Also, the leachates ob-tained fromwaste piles in the Almadén area produced total mercuryleachate concentrations ranging from 2.2 to 1400 μg/L in unfilteredsamples and from 2.0 to 400 μg/L in filtered samples (Gray et al., 2004).

Otherwise, themajor differences inmobilizedmercury in theminingwastes sample of theValle del Azoguemine (Fig. 8), could be associated,partially, with the presence of colloidal mercury. Thus, the higherconcentrations detected both in the filtered and unfiltered samplesmaybe caused by the presence of soluble chloride compounds (perhapscalomel: Hg2Cl2 and kuzminite) and possible finely particulate Hg-bearing materials or colloids released during the column experiment.Other laboratory column experiments conducted on calcines showedthe mobilization of mercury-bearing (N2 μg/g) colloids in the 50–400 nm size range associated with crystalline particles of alunite–jarosite, hematite and amorphous silica–aluminum particles (Shawet al., 2001; Rytuba, 2003; Lowry et al., 2004). Moreover, the dissolutionof secondary phase minerals, such as jarosite, natrojarosite and Fe

oxyhydroxides, could explain a percentage of the mobilized Hg if it wasscavenged in these minerals by sorption processes.

4.4. Geochemical modeling

First equilibrium speciation modeling of Hg was performed usingthe PHREEQC numerical code (Parkhurst and Appelo, 1999), howeverthe results showed that the dominant inorganic species is Hg0,including the leaching data of sample A06, which may contentsignificative amounts of Hg soluble compounds. Thus, in contami-nated groundwater under pH–Eh conditions similar to those of thecolumn experiments, the Hg speciation results showed that most ofthe Hg (84%) existed as HgCl2 and that only 4% was Hg0 (Bollen et al.,2008). Therefore, was used the numerical code MINTEQ (Allison et al.,1991), which produced more coherent results.

In the modeling of Hg speciation was considered that thegeochemical system is in equilibriumwithHg0, usingfive representativeleachates (AZ2, CAL1, AG1, A06-6A and BA1) and the pH–Eh conditionsof the column experiments (Tables 4 and 5). The results of thegeochemical modeling showed a significative variation in the distribu-tion of Hg species (Table 6). Thus, the dominant species in calcineleaching (samples AZ2 and CAL1) and Bayarqueminingwastes leaching(sample BA1)wereHg0 andHg(OH)2,while the dominant species in theminingwastes of the Valle del Azoguemine (samples AG1 and A06-6A)were HgCl2, HgCl3− and HgClOH, in concordance with the high Hg andchloride concentrations of these leachates (Table 6). These results arelike to mercury speciation analyses in groundwater, found by Bollenet al. (2008),which studied theHgspecies in groundwaterwithvery lowamounts of DOC (dissolved organic matter).

Fig. 9. Evolution of chloride, sulfate and bicarbonate from sample A06 (Valle del Azoguemining-wastes lixiviation).

Table 6Distribution of Hg species for leachates.

Species AZ2 CAL1 AG1 A06-6A BA1

HgHg0 1.26·10−7 1.26·10−7 1.26·10−7 1.21·10−7 1.26·10−7

Hg(OH)2 1.15·10−7 2.50·10−8 2.68·10−7 1.77·10−9 1.57·10−9

HgClOH 1.22·10−8 2.05·10−9 7.48·10−7 1.23·10−7 8.41·10−11

HgCl2 2.17·10−10 3.34·10−11 4.18·10−7 1.70·10−6 8.99·10−13

HgOHCO3− 2.46·10−9 5.29·10−10 5.54·10−9 3.31·10−11 8.41·10−11

HgCl3− 8.39·10−13 1.75·10−13 7.54·10−8 1.54·10−6 4.47·10−15

Data calculated using MINTEQ. Values in molality.AZ2: calcine-leachate filtered (solid sample M03), CAL1: calcine-leachate unfiltered(solid sample M03), AG1: mining wastes-leachate filtered (solid sample A06), A06-6A:mining wastes-leachate unfiltered (solid sample A06), BA1: Bayarque mining wastes-leachate filtered (solid sample B03).

Fig. 10. Mercury pH–Eh diagram.

Table 7Calculated saturation index for leachates.

Phase mineral AZ2 CAL1 AG1 A06-6A BA1

Hg mineralsCalomel (Hg2Cl2) −8.784 −11.888 −5.305 −11.553 −13.842Hg(OH)2 −3.43 −4.095 −3.065 −5.244 −5.296Montroydite (HgO) −3.312 −3.977 −2.947 −5.137 −5.178

Saturation indices calculated using MINTEQ. AZ2: calcine-leachate filtered, CAL1:calcine-leachate unfiltered, AG1: mining wastes-leachate filtered, A06-6A: miningwastes-leachate filtered, BA1: Bayarque mining wastes-leachate filtered.

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The stability of Hg species depends on the pH–Eh conditions of thecolumn experiments (Tables 4 and 5). The pH–Eh diagram indicatesthat the inorganic Hg species Hg0 may, theoretically, be the dominantinorganic Hg species in the leachates (Fig. 10). However, theinaccuracies in measuring Eh and pH could produce a deviation ofreal thermodynamic equilibrium conditions of leaching experiments.

The possible dissolution reactions, that may explain the Hgmobilization in the leaching experiments, may be suggested by thesaturation indices of the main mineral phases calculated by MINTEQnumerical code (Table 7). So, in all the leachates calomel, Hg(OH)2 andmontroydite are subsaturated, which suggests the possibility of theirdissolution.Moreover, the solubility of calomel is approximately 2mg/L(Davis et al., 1997), a concentration similar to the total Hg detected insome leachates.

5. Conclusions

The study of mine wastes (calcines and mining wastes) in the oldmining sites of Valle del Azogue and Bayarque (Almería, SE Spain)showed that the dominant mercury phases in mining wastes arecinnabar and elemental mercury bound to the solidmatrix. In calcines,however, the dominant mercury phase is only elemental mercuryassociated towastematrix. The SEM-EDS study of minewaste samplesshowed the presence of cinnabar and particles that may be associatedto calomel and kuzminite in the mining wastes sample (A06) of theValle del Azogue mine.

Water-leaching experiments were conducted on mining wastesand calcines in order to simulate the mobilization of mercury byrunoff in environmental conditions. The laboratory column-leachingexperiments indicated that the mobilization of mercury is verydifferent in calcines and mining wastes enriched in possible Hg-soluble phases. The mercury concentrations obtained from calcineleachates are very similar in the filtered and unfiltered samples,indicating the possible solubilization of Hg0, possibly readsorbed inthematrix from furnace emissions. Themajor differences in mobilizedmercury in the mining wastes samples could be associated with thepresence of colloidal mercury and Hg-soluble phases. In the leachingof mining wastes from the Valle del Azogue mine, the higherconcentrations detected in the filtered and unfiltered samples couldbe caused by the presence of soluble chloride compounds (probablycalomel and kuzminite) and finely particulate Hg-bearingmaterials orcolloids released during the column experiment.

The mercury speciation, evaluated using the MINTEQ numericalcode, showed that the theoretical dominant inorganic species in thecalcine-leaching experiments were Hg0 and Hg(OH)2. In the leachatesobtained from the Valle del Azogue mining wastes (sample A06), thedominant species were HgCl2, HgCl3− and HgClOH. These resultsdemonstrate the environmental mobility of mercury in semi-aridregions, especially during wet episodes associated with intenseprecipitation events, when the mine wastes present significativeamounts of Hg-soluble phases, like chloride compounds.

Acknowledgements

This work was supported by the Spanish Ministry of Science andTechnology (projects REN2003-09247-C04-03 and ENE2006-13267-C05-03) in collaboration with the Center for Energy, Environment andTechnology Research (CIEMAT). We thank to the anonymous reviewerstheir constructive comments that helped to improve the paper.

References

Allison, J.D., Brown, D.S., Novo-Gradac, K.J., 1991. MINTEQA2/PRODEFA2, A GeochemicalAssessment Model for Environmental Systems: Version 3.0 Users Manual. U.S.Environmental Protection Agency, Athens, GA. EPA/600/3-91/021.

APHA, 1989. Standard methods for the examination of water and wastewater, 17th ed.APHA, Washington.

Becker, G.F., 1888. Geology of the quicksilver deposits of the Pacific Slope with Atlas. USGeological Survey Library, Document 23442, Spanish localities, pp. 27–32.

Biester, H., Schulz, C., 1997. Determination of mercury phases in contaminated soils —

Hg-pyrolysis versus sequential extractions. Environment Science Technology 31,233–239.

Biester, H., Hess, A., Müller, G., 1997. Investigations on different mercury-phases in soilsof a Hg-mining area by a temperature-controlled-pyrolysis technique. In: Reuther,R. (Ed.), Geochemical Approaches for the Environmental Engineering of Metals.Springer.

Biester, H., Gosar, M., Müller, G., 1999. Mercury speciation in tailings of the Idrijamercury mine. Journal of Geochemical Exploration 65, 195–204.

Biester, H., Gosar, M., Covelli, S., 2000. Mercury speciation in sediments affected bydumped mining residues in the drainage area of the Idrija mercury mine, Slovenia.Environmental Science & Technology 34, 3330–3336.

Bollen, A., Wenke, A., Biester, H., 2008. Mercury speciation analyses in HgCl2-contaminated soils and groundwater-implications for risk assessment andremediation strategies. Water Research 42, 91–100.

Booth-Rea, G., Azañón, J.M., García-Dueñas, V., Augier, R., 2003. Uppermost Tortonian toQuaternary depocentre migration related with segmentation of the strike-slipPalomares Fault-Zone, Vera Basin (SE Spain). Comptes Rendus Geoscience 335,751–761.

Churchill, R.C., Meafhrel, C.E., Suter, P.J., 2004. A retrospective assessment of goldminingin the Reedy Creek sub-catchment, northeast Victoria, Australia: residual mercurycontamination 100 years later. Environmental Pollution 132 (2), 355–363.

Davis, A., Bloom, N.S., Hee, S.S., 1997. The environmental geochemistry andbioaccessibility of mercury in soils and sediments: a review. Risk Analysis 17 (5),557–569.

Dini, A., Benvenuti, M., Costagliola, P., Lattanzi, P., 2001. Mercury deposits inmetamorphic settings: the example of Levigliani and Ripa mines, Apuane Alps(Tuscany, Italy). Ore Geology Reviews 18, 149–167.

EPA, 1991. Description and sampling of contaminated soils. A Field Pocket Guide. EPA/625/12-91/002.

Fernández-Martínez, R., Loredo, J., Ordóñez, A., Rucandio, M.I., 2005. Physicochemicalcharacterization and mercury speciation of particle-size soil fractions from anabandoned mining area in Mieres, Asturias (Spain). Environmental Pollution 142,217–226.

Gosar, H., Pirc, S., Bidvoc, M., 1997. Mercury in the Idrija river sediments as a reflection ofmining and smelting activities of the Idrija mercury mine. Journal of GeochemicalExploration 58, 125–131.

Gray, J.E., 2003. Leaching, transport, and methylation of mercury in and aroundabandoned mercury mines in the Humboldt River basin and surrounding areas,Nevada. US Geological Survey Bulletin 2210–C.

Gray, J.E., Theodorakos, P.M., Bailey, E.A., Turner, R.R., 2000. Distribution, speciation andtransport of mercury in stream-sediment, stream-water, and fish collected nearabandoned mercury mines in SW Alaska, USA. The Science of Total Environment260, 21–33.

Gray, J.E., Crock, J.G., Fey, D.L., 2002. Environmental geochemistry of abandonedmercury mines inWest-Central Nevada, USA. Applied Geochemistry 17, 1069–1079.

Gray, J.E., Hines, M.E., Higueras, P.L., Adatto, J., Lasorsa, B., 2004. Mercury speciation andmicrobial transformations in mine wastes, stream sediments and surface waters atthe Almadén Mining District, Spain. Environmental Science & Technology 38,4285–4292.

Gustin, M.S., Biester, H., Kim, C.S., 2002. Investigation of the light-enhanced emission ofmercury from naturally enriched susbtrates. Atmospheric Environment 36,3241–3254.

Higueras, P., Oyarzun, R., Biester, H., Lillo, J., Lorenzo, S., 2003. A first insight intomercury distribution and speciation in soils from the Almadén mining district,Spain. Journal of Geochemical Exploration 80, 95–104.

Higueras, P., Oyarzun, R., Lillo, J., Sánchez-Hernández, J.C., Molina, J.A., Esbri, J.M.,Lorenzo, S., 2006. The Almadén district (Spain: anatomy of one of the world'slargest Hg-contaminated sites. Science of the Total Environment 356, 112–124.

Lechler, P.J., 1999. Modern mercury contamination from historic amalgamation millingof silver-gold ores in the Carson River, Nevada and Jordan Creek, Idaho: importanceof speciation analysis in understanding the source, mobility and fate of pollutedmaterials. In: Ebinghaus, R., Turner, R.R., de Lacerda, L.D., Vasiliev, O., Salomons, W.(Eds.), Mercury Contaminated Sites. Springer.

Loredo, J., Pereira, A., Ordóñez, A., 2003. Untreated abandonedmercurymining works ina scenic area of Asturias (Spain). Environmental International 29, 481–491.

Lowry, G.U., Shaw, S., Kim, C.S., Rytuba, J.J., Brown, G.E., 2004. Macroscopic andmicroscopic observations of particle-facilitated mercury transport from New Idriaand sulphur bank mercury mine tailings. Environmental Science & Technology 38,5101–5111.

Martinez, J., Navarro, A., Lunar, R., 1997. First reference of pyrite framboids in a Hg–Sbmineralization: the Valle del Azogue mineral deposit (SE Spain). Neues JahrbuchMineral Mondtsh, Jg. 4, 175–184.

Mendoza, J.L., Navarro, A., Cuitiño, L., 2006. Nueva interpretación metalogenética delyacimiento de mercurio del Valle del Azogue (Almería), España. XI CongresoGeológico Chileno, Actas, vol. 2.

Morel, F.M.M., Kraepiel, A.M.L., Amyot, M., 1998. The chemical cycle and bioaccumula-tion of Mercury. Annual Review of Ecology and Systematics 29, 543–566.

Navarro, A., Martínez, J., Font, X., Viladevall, M., 2000. Modelling of modern mercuryvapor transport in an ancient hydrothermal system: environmental and geochem-ical implications. Applied Geochemistry 15 (3), 281–294.

Navarro, A., Biester, H., Mendoza, J.L., Cardellach, E., 2006. Mercury speciation andmobilization in contaminated soils of the Valle del Azogue Hg mine (SE, Spain).Environmental Geology 49, 1089–1101.

245A. Navarro et al. / Journal of Geochemical Exploration 101 (2009) 236–246

Author's personal copy

Parkhurst, D., Appelo, C., 1999. User's guide to PHREEQC (Version2). A computer programfor speciation, batch-reaction, one-dimensional transport, and inverse geochemicalcalculations. Water-Resources Investigations Report, vol. 99–4259. U.S. GeologicalSurvey.

Puigdomenech, I., 2004. Make Equilibrium using Sophisticated Algoritms (MEDUSA)program, Inorganic Chemistry Department. Royal Institute of Technology, 100 44,Stockholm, Sweeden. http://web.telia.com/.

Qiu, G., Feng, X., Wang, S., Shang, L., 2005. Mercury and methiylmercury in riparian soil,sediments, mine-waste calcines, and moss from abandoned Hg mines in eastGuizhou province, southwestern China. Applied Geochemistry 20, 627–638.

Ravichandran, M., Aiken, G.R., Reddy, M.M., Ryan, J.N., 1998. Enhanced dissolution ofCinnabar (Mercuric Sulfide) by dissolved organic matter from the FloridaEverglades. Environmental Science & Technology 32, 3305–3311.

Ravichandran, M., Aiken, G.R., Reddy, M.M., Ryan, J.N., 2003. Enhanced dissolution ofCinnabar (Mercuric Sulfide) by dissolved organic matter from the Florida Everglades.USGS WRDWeb Page.

Rytuba, J.J., 2000. Mercury mine drainage and processes that control its environmentalimpact. The Science of Total Environment 260, 57–71.

Rytuba, J.J., 2002. Mercury geoenvironmental models. In: Seal, R.R., Foley, N.K. (Eds.),Progress on Geoenvironmental Models for Selected Mineral Deposit Types. USGeological Survey. Open-File Report 02-195.

Rytuba, J.J., 2003. Mercury from mineral deposits and potential environmental impact.Environmental Geology 43, 326–338.

Rytuba, J.J., 2005.Geogenic andmining sources ofmercury to the environment. In: Parsons,M.B., Percival, J.B. (Eds.), Mercury, Sources, Measurements, Cycles and Effects. ShortCourse Series, vol. 34. Min. Ass. of Canada.

Rytuba, J.J., Kotlyar, B.B., Wilkerson, G., Olson, J., 2001. Geochemistry of SelectedMercuryMine-tailings in the Parkfield Mercury District, California. USGS, Open-FileReport 01-336, version 1.0.

Shaw, J., Lowry, G.U., Kim, C.S., Rytuba, J.J., Brown, G.E., 2001. The influence of colloidalphases on Hg-transport from mercury mine waste tailings: a laboratory case studyof the New Idria and Sulphur Bank Mines, California, USA. Eleventh Annual V.M.Goldschmidt Conference. 3723 pdf.

Shaw, S.A., Al, T.A., MacQuarrie, K.T.B., 2006. Mercury mobility in unsaturated goldminetailings, Murray Brook mine, New Brunswick, Canada. Applied Geochemistry 21,1986–1998.

Viladevall, M., Font, X., Navarro, A., 1999. Geochemical mercury survey in the AzogueValley (Betic area, SE Spain). Journal of Geochemical Exploration 66, 27–35.

Wang, Q., Kim, D., Dionysiou, D.D., Sorial, G.A., Timberlake, D., 2004. Sources andremediation for mercury contamination in aquatic systems-a literature review.Environmental Pollution 131, 323–336.

Ying, Y., Allen, H., Li, Y., Huang, C.P., Sanders, P.I., 1996. Adsorption of mercury (II) by soil:effects of pH, chloride, and organic matter. Journal of Environmental Quality 25, 837.

Zhang, G., Liu, C.Q.,Wu, P., Yang, Y., 2004. The geochemical characteristics of mine-wastecalcines and runoff from theWanshanmine, Guizhou, China. Applied Geochemistry19, 1735–1744.

246 A. Navarro et al. / Journal of Geochemical Exploration 101 (2009) 236–246