Bacterial influence on storage and mobilisation of metals in iron-rich mine tailings from the Salobo mine, Brazil

Science of the Total Environment 680 (2019) 91–104

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Bacterial influence on storage and mobilisation of metals in iron-rich mine tailings from the Salobo mine, Brazil A. Henne a,⁎, D. Craw b, E.J. Gagen a, G. Southam a a b

School of Earth and Environmental Sciences, The University of Queensland, St Lucia, QLD 4072, Australia Department of Geology, The University of Otago, North Dunedin, Dunedin 9016, New Zealand

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• Fe(II)-silicates in tailings can sustain A. ferrooxidans for ~400 days. • Circumneutral pH prevails in Fe(II)-silicate rich tails dominated by A. ferrooxidans. • Most trace metals in solution are adsorbed onto biogenic ferrihydrite. • A. ferrooxidans enhance silicate dissolution reactions that control water composition. • Without acid addition, the long-term pH of tailings will remain circumneutral.

a r t i c l e

i n f o

Article history: Received 5 March 2019 Received in revised form 30 April 2019 Accepted 30 April 2019 Available online 6 May 2019 Editor: Frederic Coulon Keywords: Copper Tailings Bioleaching Salobo Acidithiobacillus ferrooxidans Metal mobility

⁎ Corresponding author. E-mail address: [email protected] (A. Henne).

https://doi.org/10.1016/j.scitotenv.2019.04.448 0048-9697/© 2019 Elsevier B.V. All rights reserved.

a b s t r a c t In this study we investigated the potential effects of promoting bacterial activity on tailings from the Salobo iron-oxide copper‑gold (IOCG) mine, Brazil. In particular we focussed on (1) the potential for mobilising additional Cu and (2) the effects of long-term storage on other metals. Unlike typical sulphide-ore tailings, the pH of the Salobo tailings is circumneutral and these tailings are dominated by Fe-bearing silicates and magnetite, with minor micrometre-scale encapsulated Cu-bearing sulphides. While these tailings do not produce acid mine drainage, an endemic strain of Acidithiobacillus ferrooxidans was isolated from the mine site. These bacteria were used in laboratory column leaching experiments of tailings material, which ran for up to 395 days, without the addition of ferrous iron. Bacteria-tailings interactions were typically maintained at a pH N 5, due to silicate-mediated pH buffering. This was eventually overcome after ~200 days by regular addition of acidic (pH 2.2) nutrient solution, as well as growth and acid generation by bacteria. Copper dissolution was not significantly enhanced by bacterial activity compared to abiotic control experiments while pH was N5. However, as the experiments were progressively acidified, additional Cu was mobilised in the biotic systems. The mineral alteration reactions produced abundant ferrihydrite precipitates within the tailings, which was enhanced by bacterial activity as the pH decreased. Adsorption of metal cations to these precipitates ensured that effluent solutions had only low levels (b0.5 mg/l) of dissolved trace metals such as As, Co, Pb, Zn, Se, Ni and Cr. These adsorption processes will strongly inhibit metal leaching from the tailings during long-term storage, as long as the iron oxidising bacteria are

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producing the requisite excess of ferrihydrite and the pH is N5. This case study shows that bacteriallymediated silicate weathering, in Fe(II)-bearing silicate rich tailings with only minor sulphides and Acidithiobacillus ferrooxidans can enhance the environmental stability of the tailings. © 2019 Elsevier B.V. All rights reserved.

1. Introduction Mine tailings are voluminous waste products of processing and metal extraction at most mines, and these tailings are typically stored on mine sites in perpetuity (Younger et al., 2002; Dold, 2008; Lottermoser, 2010). These tailings represent large potential future resources, which have already been mined and crushed, and therefore are readily available for extraction of additional metals when technology and economic conditions permit (Dold, 2008; Lottermoser, 2010). Until such time, it is important that the tailings be stored in an environmentally stable form, with strict limitations on undesirable solid and dissolved discharges (Dold, 2008; Lottermoser, 2010). Bacterial processes within the tailings can cause acidification and mobilise metals into interstitial solutions and discharging waters (Bhatti et al., 1994; Baker and Banfield, 2003; Southam and Saunders, 2006; Diaby et al., 2007; Diaby et al., 2015). The role of iron- and sulphur-oxidising bacteria in the formation of acid mine drainage (AMD) (Baker and Banfield, 2003; Johnson and Hallberg, 2003; Dold, 2014; Quatrini and Johnson, 2018) and the role of microorganisms such as sulphur-reducing bacteria in the remediation of AMD (Johnson and Hallberg, 2005; Johnson, 2014; Kefeni et al., 2017) is well studied. The use of these processes for the extraction of metals from mine waste material can be applied on an industrial scale in bioleaching operations (Brierley and Brierley, 2013; Watling, 2015).

(a)

However, relatively little is known about the interaction of bacteria and minerals in alkaline or neutral tailings. Most recently, work has focussed on bacterial communities from alkaline (Sun et al., 2018; Li et al., 2014) or moderately acidic (Mendez et al., 2008) tailings for soil development during tailings revegetation, and only a handful of studies examine the evolution of microbiomes from the initially neutral or moderately acidic stages of sulphidic tailings to the highly oxidised stages (Huang et al., 2011; Chen et al., 2013; Liu et al., 2014; Korehi et al., 2014). From a conventional geochemical perspective neutral tailings with little sulphide contents do not pose a liability for AMD development. However, at neutral pH hydrogen ions are still released during ferric hydroxide formation (Johnson and Hallberg, 2005; Taylor et al., 2005) and some metal ions can remain in solution (Moodley et al., 2018). Furthermore, the activity of moderately acidophilic bacteria can alter the geochemistry of initially moderately acidic tailings to the point where acidophilic iron- and sulphur-oxidising bacteria become the dominant species at lower pH (Korehi et al., 2014). Therefore, interactions between bacteria and alkaline tailings can be both environmentally and economically significant. In this study, we investigate the interactions between bacteria and alkaline (pH 7.0–8.2) tailings derived from the Salobo iron-oxide copper‑gold (IOCG) mine in Brazil (Fig. 1a–d). The mine tailings are dominated by ferrous iron-bearing silicates and oxides. Minor Cu-

(b) Venezuela Colombia

Guyana Suriname French Guiana

Salobo

North Atlantic Ocen

Ecuador

National Forest*

Salobo Peru

BRAZIL

South Pacific Ocean

20 km

Bolivia

(c)

Tailings Dam

(d)

Processing Plant

Pit 1 km

Fig. 1. (a) Location of Salobo in Brazil, South America (ArcMap 10.3). (b) Location of Salobo mine within the Tapirapé National Forest in Brazil (GoogleEarth). *The forested area includes the National Forests Tapirapé-Aquiri, Carajás and Itacaiúnas, the Tapirapé Biological Reserve, Igarapé Gelado EPA, and Xikrin-Cateté Indigenous Land. (c) Arial view of Salobo mine pit, processing plant and tailings dam (GoogleEarth). (d) View over parts of the Salobo mine tailings in April 2018.

A. Henne et al. / Science of the Total Environment 680 (2019) 91–104

environments (Henne et al., 2018), which is unusual compared to the acidic environments preferred by this organism (pH b3; Amaro et al., 1991). The Salobo mine site lies within the Tapirapé National Forest in the Amazon rain forest of Brazil (Fig. 1a–d), and is therefore particularly environmentally sensitive with respect to long-term tailings storage and potential acidification and metal discharge. As one of the largest Cu deposits in Brazil, the tailings dam is predicted to hold up to 565 Mm3 of material over an area of 13 km2 (Vos et al., 2015). The tailings material examined in this study typically contains 0.1% Cu (Table 1) that is not extracted by current processing methods. Therefore, the long-term resource potential of the tailings is substantial, with between 105 and

bearing sulphides are typically encapsulated in these silicates and oxides and therefore bypass the mine concentration system (Fig. 2a–e). Because of the overwhelming abundance of Fe (II)-bearing minerals within the tailings, this study focuses on a characterised and effective iron oxidising bacterium, A. ferrooxidans, which is common in global surficial mine environments (Baker and Banfield, 2003), and widely used in engineered bioleaching operations (Bosecker, 1997; Brierley and Brierley, 2013; Watling, 2015). An endemic strain of this bacterium was previously cultured from the Salobo mine site (Henne et al., 2018) to investigate leaching of Fe(II)-bearing material from Salobo. This endemic bacterial strain is notable for its ability to survive at the site by oxidising Fe(II)-bearing silicates in ambient circumneutral pH

(a)

93

Copper Minerals 0.1 %

(c) Chlorite 7% Grunerite 8% Biotite 30 %

Other* 9%

1mm Magnetite 12 %

(b) Garnet 22 %

Biotite

Garnet

Grunerite

Chlorite

Quartz

Magnetite

1mm

(d)

Quartz 12 %

(e)

10µm

10µm

Fig. 2. Mineral composition and texture of Salobo tailings prior to leaching. (a) Backscatter image of a polished block of pulverised tailings material in columns prior to leaching. (b) Mineral Liberation Analysis map of (a). (c) Major phases in tailings material. Arrows indicate Fe(II) potentially available for biooxidation. *Ca-amphibole, plagioclase, allanite, orthoclase, olivine, Ca-pyroxene, ilmenite, apatite, muscovite, calcite. (d) Backscatter image of Cu-sulphides encapsulated in biotite. (e) Backscatter image of Pb-bearing Cu-sulphides in amphibole.

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Table 1 Element concentrations in water from the Salobo River and the tailings dam, and in pulverised tailings material. All values are reported in ppm, except where indicated otherwise. Location K4 from Konhauser et al. (1994) is approximately at the current mine site. Konhauser et al. (1994) location K2 is upstream and K10 downstream from the current mine site. Konhauser et al. (1994)⁎

Al As B Ba Ca Cd Co Cr Cu Fe K Mg Mn Mo Na Ni P Pb S Se Zn TOC TC TIC TN pH

K2

K4

K10

0.1 bd.l. 0.2 0.1 2.1 bd.l. 0.0 0.0 bd.l. 0.4 2.1 1.9 0.0 0.0

0.0 bd.l. 0.0 0.0 2.4 bd.l. 0.0 0.0 bd.l. 0.3 1.2 2.4 0.0 0.0

0.1 bd.l. 0.0 0.1 2.2 bd.l. 0.0 0.0 bd.l. 0.7 1.2 2.4 0.0 0.0

0.0 bd.l. 0.0

0.0 bd.l. 0.0

0.0 bd.l. 0.0

bd.l. 0.0

bd.l. 0.0

bd.l. 0.1

Damous et al. (2002) Dry season

Wet season

6.5

6.2

0.0

0.0

0.9 0.1

1.1 0.1

10.7

10.5

0.3

0.0

3.1 7.8^ 4.6 0.9 6.8

2.4 0.3 6.9

Salobo Tailings 2018⁎

Pulverised Tailings 2016

AH 1811D

Column Material

0.1 bd.l. 0.1 0.1 14.7 bd.l. bd.l. bd.l. 0.1 0.5 6.9 3.8 0.6 bd.l. 9.6 bd.l. bd.l. bd.l. 1.8 bd.l. bd.l. 4.0 17.9 13.9 2.6 7-7.4

4.8 % bd.l. nd 185.7 mg/kg 1.1 % 0.1 mg/kg 99.7 mg/kg 170.0 mg/kg 0.1 % 30.2 % 1.3 % 1.4 % 0.6 % nd 0.4 % 98.3 mg/kg bd.l. 11.7 mg/kg 0.4 mg/kg nd 34.9 mg/kg nd 0.2% nd b0.5 % 8.11⁎⁎

bd.l. below detection limit; TOC = total organic carbon; TC = total carbon; TIC = total inorganic carbon; TN = total nitrogen; nd = not determined ⁎ Dry season only. ⁎⁎ paste pH. ^ Calculated value.

106 t of Cu potentially available for future extraction. Hence, we used laboratory leaching column experiments, without the addition of iron, to define the likely long-term pH trajectory and levels of metal mobilisation, including Cu, that will arise within the tailings storage facilities at the mine. In this study we address both the potential environmental effects of interactions between endemic A. ferrooxidans and the tailings minerals in the context of long-term storage, and the potential use of this endemic strain for extraction of additional Cu from the tailings. 2. Materials and methods 2.1. Mine site and material characterisation The Salobo mine is being developed in an Archean iron oxide copper gold (IOCG) deposit (Requia and Fontboté, 1999; Requia and Fontboté, 2000; PorterGeo, 2001; Requia et al., 2003; Williams et al., 2005; Grainger et al., 2008; Moreto et al., 2015). The 100–600 m wide mineralisation zone extends ~4 km along strike with average Cu contents of bulk ore ranging from 0.6% to 1.1% with 10% to 50% magnetite contents (PorterGeo, 2001; Gurmendi, 2004). The mineralisation is hosted in Fe-rich schists and is mined as an open pit operation (Burns et al., 2017). The sulphide ore contains large amounts of iron oxides, mainly magnetite (Fe3O4), with bornite (Cu5FeS4), chalcocite (Cu2S) and minor chalcopyrite (CuFeS2) (Requia and Fontboté, 1999; Requia and Fontboté, 2000; PorterGeo, 2001; Burns et al., 2017), and variable proportions of amphibole, olivine, garnet, biotite, quartz and plagioclase (PorterGeo, 2001). Ore is crushed to a 20 μm particle size to liberate

sulphide minerals. The sulphides are then concentrated by flotation and lime is added during processing (Burns et al., 2017). Mine tailings consist mainly of Fe-rich silicate minerals with minor remnant sulphide grains encapsulated in the silicate particles. Samples for this study for DNA extraction, pH and water analyses were collected from the tailings impoundment at the Salobo mine during the dry season in April 2018. Samples of bacteria for inoculation of leaching columns were collected from fractures in the active Salobo mine in February 2016 (Henne et al., 2018). Material for the leaching columns was obtained from Salobo mine staff and consists of dried and pulverised material from the flotation plant, intercepted before it entered the tailings impoundment. Major elements and Cu in tailings material were analysed by inductively coupled plasma optical emission spectrometry (ICP-OES) on an Optima 8300DV after fusion. Trace elements were analysed by inductively coupled plasma mass spectrometry (ICP-MS) using a 7900 Agilent after digestion following the ICP-MS Collision/Reaction Cell Procedure. X-ray diffraction (XRD) analysis was carried out on the powdered material using a Bruker D8 Advance instrument. Mineral identifications were validated against mineral assemblages previously reported for the Salobo deposit (Lindenmayer, 1990; Veiga et al., 1991; Requia and Fontboté, 1999; Requia and Fontboté, 2000; Requia et al., 2003; Tassinari et al., 2003; deMelo et al., 2016; Henne et al., 2018). Mineral liberation analysis (MLA) of the pulverised material was carried out to quantify the mineral proportions, their particle sizes, and provide indications of exposed surfaces. The material was split in a rotary sample riffle and embedded in Struer Epofix resin. The surface of the resin block was polished to 1 μm and coated with carbon. Sample preparation and measurements were carried out using an FEI Quanta 600 MK2 Scanning Electron Microscope with 1.16 μm per pixel. Measurements of pH were completed in the field on running surface water within the tailings impoundments using pH strips, and the pH of four tailings samples, as well as the pulverised material, was measured using the paste pH method for soil and waste pH in the laboratory (Method 9045D, US EPA). Water samples from the tailings impoundment were analysed for major and trace elements by ICP-OES. Elemental microanalysis for carbon and nitrogen was carried out on a FLASH 2000 CHNS/O Analyser following a modification of the Pregl and Dumas method. 2.2. Bacterial leaching columns Bacterial enrichments for inoculation of the leaching columns came from an acidophilic iron oxidising enrichment culture, grown from a sample collected from a fracture with seeping ground water at the Salobo mine pit (Henne et al., 2018). Enrichment cultures were grown in a nutrient solution modified after Silverman and Lundgren (1959) containing 0.4 g/l (NH4)2SO4, 0.1 g/l K2HPO4 and 0.4 g/l MgSO4·7H2O and 33.3 g/100 ml FeSO4·7H2O. Viable cell counts were determined using the Most Probable Number (MPN) statistical method modified after Cochran (1950) using the above medium. Leaching columns were prepared in triplicate for bacterial leaching, plus one abiotic control column. Ten grams of tailings material and ca. 1 cc of glass wool were loaded into 10 cc sterile, polypropylene syringes. One biotic column did not contain glass wool to prevent loss of powdered material from the column, as other filtering methods were trialled but had to be abandoned after sample loss. All loaded columns were gamma irradiated at 50 kGy for sterilisation. Biotic columns were inoculated with ca. 5 × 107 cells/ml in 5 ml of nutrient solution (as above, but without the addition of Fe), and incubated for one week at 30 °C. The abiotic control column was ‘inoculated’ with sterile nutrient solution for 24 h. Sodium benzoate (0.3 g/l; modified after Norris et al. (2012) and Watling et al. (2009)) was added to the abiotic control column to inhibit growth of acidophilic bacteria and archaea. Columns were irrigated with nutrient solution every second to third day at a rate of 1 ml/day. The dosage of sodium benzoate in the control column

A. Henne et al. / Science of the Total Environment 680 (2019) 91–104

OTU 1 2 3 4 6 5 7 11 14 12 10 08 09 13 15 17 16 21 20 27 18 22 39 25 23 24 26 19 28 30

Biol. Column Column Inoculum #2

Biol. Column #3

96.1 %

93.3 %

90.4 %

Not detected

Tailings Sample A

Tailings Sample B 59.7 % 13.8 %

14.1 % 5.8 % 5.4 % 5.0 %

95

Nearest named isolate and BLAST identity

100% 100% 98% 100% 100% 100% 100% 100% 99% 100% 100% 100% 100% 94% 94% 92% 100% 100% 100% 100% 100% 100% 92% 100% 100% 99% 100% 99% 100% 98%

Acidithiobacillus ferrooxidans MK123861.1 Serratia marcescens MK414953.1 Trichocoleus desertorum EU586742.1 Thiobacillus thioparus HM535225.1 Algoriphagus aquatilis MK039096.1 Stenotrophomonas maltophilia MK641667.1 Hydrogenophaga bisanensis MG712817.1 Beta proteobacterium BIWA26 LC217413.1 Ideonella sp. str. CUAQ17 MK503699.1 Rhodocyclaceae bacterium ICHIDE20 LC132836.1 Novosphingobium aromaticivorans MK402961.1 Enterobacter sp. SA187 CP019113.1 Microcella sp. RO61 HE863752.1 Bacterium Ellin5280 AY234631.1 Thiobacillus sp. AF023264.1 Gemmatirosa kalamazoonesis CP007128.1 Mariniradius saccharolyticus NR_117078.1 Flavihumibacter cheonanensis LN867289.1 Dechloromonas agitata MG205615.1 Bacillus sp. str. DW024 MK713577.1 Lysobacter sediminicola NR_147745.1 Lacunisphaera limnophila NR_146349.1 Methanosaeta pelagica NR_113571.1 Sediminibacterium salmoneum NR_044197.1 Beta proteobacterium HTCC349 AY429717.1 Rhodobacter sp. HME865 KC157045.1 Beta proteobacterium WD179 HQ341760.1 Nostoc sp. MACC-231 MH702233.1 Phreatobacter oligotrophus NR_133817.1 Aquicella siphonis AY359284.1

OTU abundance

Fig. 3. Heatmap analysis of 16S rRNA gene sequences of experiment inoculum, leaching column samples at 185 days, and natural samples showing the 30 most abundant OTUs (operational taxonomic units), when sequences were clustered at a distance of ≤0.03. The nearest named isolate in the public domain and its accession number is given for each OTU. OTUs highlighted in blue demonstrate b95% 16S rRNA gene identity to a cultured isolate. The scale bar relates to the relative abundance of OTUs within the respective samples, ranging from black (least) to red (most) abundant. White spaces indicate undetected OTUs. The abundance of OTUs that are at N5% abundance within a library have been indicated as a proportion within each column.

nutrient solution was increased to 0.6 g/l on day 53 to maintain sterile conditions after observation of a single bacterium in the leach solution. In hindsight, the observation of a single, killed bacterium over the entire experiment, from natural materials known to possess bacteria should not have been surprising. Column effluents were collected every two days (=2 ml), stored below −5 °C, and combined to 10 day (=10 ml) batches for analysis. The pH of effluents was tested every 10 days using a pH 700/ion benchtop meter, and free bacteria counts were carried out using a Petroff Hausser counting chamber and a Nikon Eclipse E200 phase contrast light microscope. The effluents were filtered using 0.45 μm syringe filters before major and trace element analysis by ICP-OES. Copper was analysed using a 0.2 ppm, 0.5 ppm, 1 ppm, 5 ppm, 25 ppm, 50 ppm and 100 ppm standard curve. Evaporation losses within the incubator were determined by adding 10 ml each of water to 5 unsealed Sarstedt tubes. Tubes were weighed before the trial and after 10 days in the incubating chamber. All ICP-OES results were thence adjusted by an average factor of 3.95% to account for evaporation losses. To meet quarantine standards, sealed effluents were heated for a minimum of 30 min at 101 °C on a heating block before they were acidified and analysed by ICP-OES. Results of effluent analyses from inoculated columns are reported as averages of triplicates with standard deviation (n = 3) for days 30 to 180. Due to sample loss n = 2 for days 0, 10 and 20. The original

experiment was designed for a 185 day period, at which point two biotic columns and the control column were harvested for further analyses. However, one biotic column was kept running to confirm trends observed during the experiment and samples were analysed in 3 × 40 day (=40 ml) batches from day 270 to 390. DNA extraction was performed on collected tailings material, on the inoculum for leaching columns and on post leach column material (including abiotic controls) as per Qiagen DNeasy Powersoil Kit (cat #12888-100). DNA extraction was followed by polymerase chain reaction (PCR) and amplification of the V6-V8 region of the 16S rRNA gene using primers 926F and 1392R (see Gagen et al. (2018) for primer sequences and details of library preparation) and sequencing was performed at the Australian Centre for Ecogenomics, the University of Queensland. Raw data was processed using MOTHUR v 1.42.0 as per the MiSEQ SOP (https://mothur.org/wiki/MiSeq_SOP) with modifications as described by Gagen et al. (2018) except that singletons in the datasets were retained in the analyses. Sequences have been submitted to the National Center for Biotechnology Information (NCBI) Sequence Read Archive under BioProject ID PRJNA540174. Scanning electron microscopy (SEM) and energy dispersive x-ray spectroscopy (EDS) were carried out on aliquots of material after 185 and 395 days of leaching to examine colonisation of minerals by iron oxidising bacteria. Whole-mount samples were fixed with 2.5%(aq) glutaraldehyde, followed by an ethanol dehydration series (20%, 40%, 60%,

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Fig. 4. Secondary electron microscopy images of material after leaching. (a) Secondary electron image of whole mound of P and S bearing ferrihydrite precipitates in tailings material after leaching for 385 days. (b) Two A. ferrooxidans cells on ferrihydrite precipitate of (a). (c) Secondary electron image of likely trace-fossil of now-dislodged A. ferrooxidans in tailings material after leaching for 185 days. (d) Backscatter image of A. ferrooxidans fossils in a ferrihydrite coating around magnetite grains in thin section from Salobo mine waste material leaching experiments. (e) Secondary electron microscopy image of extracellular polymeric substances (EPS) on tailings material after leaching for 185 days.

80%, 3 × 100%) and critical point drying. They were mounted on stainless steel stubs using adhesive carbon tabs and carbon coated. To examine the leached material in situ a biotic column of Salobo tailings and another biotic column of mineralogically comparable crushed waste rock (described in Henne et al., 2018), were fixed with 2.5%(aq) glutaraldehyde and embedded in a polyhydroxy-aromatic acrylic resin (LR White) following a procedure modified after Tippkötter et al. (1986) and Mahrous (2012). The columns were washed with ~10 ml of DI water followed by ~5 ml of 1%(aq) osmium tetroxide, ~10 ml of ethanol at 20%, 40%, 60%, 80% and ~35 ml of 100% ethanol, before ~10 ml of 1:1100% ethanol and LR White was added to the columns. As the final step ~10 ml of LR White was allowed to flow through the columns. Flow rates of liquids added to the tailings columns were ~5 ml/day, while each liquid was kept in the syringe containing crushed waste rock for a minimum of 10 min before it was drained. The columns were placed into 50 ml falcon tubes, completely filled with LR White to exclude any oxygen, sealed, left to rest overnight and oven-dried at 60 °C for 24 h. The columns were then cut, polished and carbon coated. SEM and EDS analyses were completed using a JOEL 7001F FE-SEM. 3. Results 3.1. Salobo tailings biogeochemistry The Fe-rich (N30 wt%) tailings material consists in large parts of the Fe-bearing silicates biotite, grunerite and chlorite, as well as garnet, quartz and the iron oxide magnetite (Fig. 2a–c). Very little Fe- and Cusulphides remain in the processed tailings material (b0.1 wt%; Fig. 2c, Table 1). The main Cu-bearing phases were bornite and chalcopyrite, which frequently occurred as inclusions in silicates (Fig. 2de). Other metal(loid)s of interest, such as As, Co, Cr, Ni, Pb and Zn occurred in

trace amounts. (Table 1). These metals typically occurred in solid solution within sulphides (e.g., Pb; Fig. 2e). Running surface water within the tailings impoundment had circumneutral pH values ranging from pH 7.0–7.4, while paste pH values obtained from tailings samples were slightly higher (pH 8.0–8.2; Table 1). Analyses of water collected from the tailings impoundment indicated that it was dominated by Ca (14.7 mg/l) followed by Na (9.6 mg/l), K (6.9 mg/l) and Mg (3.8 mg/l), with relatively high concentrations of S (1.8 mg/l), Mn (0.6 mg/l) and Fe (0.5 mg/l; Table 1) for what is considered unreactive material. Copper concentrations were low (0.05 mg/l) and other trace metals were below the detection limit (Table 1). Significant nitrogen (2.6 mg/l) and dissolved total carbon concentrations (17.9 mg/l) occurred in the water (Table 1). Alpha diversity data of the two tailings samples highlights the variability in the diversity of the tailings. Tailings Sample B showed low diversity with a Shannon index of 1.8, while Tailings Sample A showed substantially higher diversity based on a Shannon index of 4.2. Major OTUs demonstrated sequence similarity to common soil bacteria (OTUs 2 and 5), a desert cyanobacterium (OTU 3), a sulphur oxidiser (OTU 4) and various heterotrophic bacteria (OTUs 6 and 7; Fig. 3). Interestingly, known acidophilic iron oxidising bacteria were not detected from either of the natural samples (Fig. 3). Rarefaction analyses (Supplementary Fig. 1) showed that there may be more diversity to be uncovered with further sequencing of Tailings Sample A. A culture grown from Salobo tailings material in 2016 produced iron oxide precipitates present as ‘bathtub’ rings on culture tubes, which indicated the presence and activity of acidophilic iron oxidising bacteria (data not shown). However, the major OTUs in this mixed culture were not closely related to any described lineages of acidophilic iron oxidisers (data not shown). It therefore seems likely that there are novel iron oxidising bacteria in the tailings and the cultures from it, and future

A. Henne et al. / Science of the Total Environment 680 (2019) 91–104

(a)

97

9 A. ferrooxidans

8

Control

pH

7 6 5 4 3 0

50

100

150

200

250

300

350

400

Days

(b) 80

(e) 1200 1000 S [mg/L]

Al [mg/L]

60 40 20

800 600 400 200

0

Feedstock

0 0

100

200

300

400

(c) 20

0

(f)

100

200

300

400

35 30 25

Cu [mg/L]

Mn [mg/L]

15 10 5

20 15 10 5 0

0 0

100

200

300

0

400

(d) 50

50

100

150

200

250

300

350

400

(g) 0.3

Pb [mg/L]

Fe [mg/L]

40 30 20

0.2

0.1

10 0

0

0

50

100

150

200

250

300

350

400

0

100

200

300

400

Fig. 5. pH and concentration of metals in biotic and abiotic column effluents over time. Aluminium, Mn and Fe concentrations are indicative of silicate dissolution. Copper, Pb and S concentrations are indicative of sulphide dissolution. (a) pH. (b) Aluminium. (c) Manganese. (d) Iron. (e) Copper. (f) Lead. (g) Sulphur.

investigation using metagenomic approaches to elucidate the organisms responsible for this phenotype would be worthwhile. 3.2. Microbial community in leaching columns Most probable number (MPN) counts on post-leach column material indicated large populations of viable acidophilic iron oxidising bacteria at N106 cells/g in inoculated columns. Orange-brown ferrihydrite precipitates, possibly with admixed nano-particulate vivianite type specific to A. ferrooxidans (Norris, 1990) were also observed on the walls of biotic columns. DNA extraction was successful for inoculated columns with A. ferrooxidans as the dominant species (N93% and N96%,

respectively; Fig. 3). Most probable number (MPN) counts of the control column confirmed that no viable Fe(II)-oxidising cells were present in the abiotic column. DNA extraction was likewise unsuccessful for this control column. Secondary electron images of post-leach material from the leaching columns showed a patchy distribution of A. ferrooxidans cells preserved in nano-particulate ferrihydrite precipitates with micron-scale elongated forms (Fig. 4a, b). Abundant EPS (extracellular polymeric substances) and occasional bacterial trace fossils were also present throughout the inoculated samples (Fig. 4c, e). Backscattered electron images of an in situ section through a 395 day biotic leaching column of waste rock material, which is mineralogically comparable to the

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Table 2 Typical effluent analyses for abiotic column effluents after 50 and 150 days and for inoculated column effluents after 50, 150 and 350 days and. All values are in mg/l. Inoculated

Control

Day

50

150

350

50

150

Al As B Ba Ca Cd Co Cr Cu Fe K Mg Mn Mo Na Ni P Pb S Se Zn

9.8 b0.2 0.3 0.1 136.8 b0.2 0.2 b0.2 0.1 0.0 80.1 73.1 12.8 b0.2 5.5 0.1 b1.0 0.0 341.5 b0.2 b0.2

45.1 b0.2 0.1 b0.2 85.9 b0.2 0.2 b0.2 17.9 0.1 58.9 65.5 7.1 b0.2 2.5 0.1 b1.0 0.1 318.0 b0.2 0.1

42.2 b0.2 0.3 b0.2 62.1 b0.2 0.2 b0.2 31.4 b1.0 58.2 64.3 6.6 b0.2 1.2 0.1 b1.0 0.1 326.0 b0.2 0.1

13.9 b0.2 0.3 0.1 183.6 b0.2 0.2 b0.2 0.2 0.1 85.0 81.6 17.0 0.1 72.5* 0.1 b1.0 0.1 393.2 b0.2 b0.2

67.8 b0.2 0.1 b0.2 122.7 b0.2 0.2 b0.2 15.9 7.9 67.0 79.3 6.4 b0.2 153.5* 0.2 b1.0 0.2 452.4 b0.2 0.1

bBelow detection limit; *sodium benzoate added to abiotic control column nutrient solution.

tailings material, but contained higher amounts of sulphur, showed numerous bacterial fossils in ferrihydrite precipitates (Fig. 4d) similar to “encrusted bacterial fossils” described in Westall (1999). 3.3. Leaching column chemistry Despite adding acidified nutrient solution (pH 2.2) at a rate of 1 ml/day, the pH in biotic leaching columns remained near pH 8 for the first 30 days of the leaching trials (Fig. 5a). The pH gradually decreased and then levelled off at pH ~5 from days 80 to 200. After 390 days, the final pH reached was pH 3.4. In comparison, the pH in the control column initially followed a similar path to the biotic columns, and then decreased steadily to pH 4 after 180 days (Fig. 5a). Dissolved Al increased in solution with time, as the pH decreased (Fig. 5b), in parallel with Mn (Fig. 5c). In the control column, dissolved Fe was low initially, then markedly increased from day 100 with concentrations of up to ~40 mg/l, coinciding with a pH drop to below pH 5 (Fig. 5d). In contrast, in the biotic columns, the dissolved Fe in solution remained largely below 0.1 mg/l (Fig. 5d). Potassium and Mg contents of effluents were elevated compared to the added nutrient solution levels of ~40 mg/l (Table 2). Concentrations of Na were low (b6 mg/l) in biotic columns and decreased with time (Table 2), while the Na concentration in the control column to which we added sodium benzoate was high to begin with (~70 mg/l) and more than doubled over time (~150 mg/l after 150 days; Table 2) with increasing addition of the bactericide. Calcium contents of effluents from biotic and abiotic columns show a general downward trend over time, which was more pronounced in biotic columns (Table 2). In contrast, dissolved P remained persistently below the detection limit in both biotic and abiotic columns despite regular addition of ~18 mg/l with the nutrient solution (Table 2). While S was added to the nutrient solution at ~150 mg/l, relatively high concentrations of dissolved S, exceeding the S addition, were measured in leach liquors: ~350 mg/l in biotic and ~400 mg/l in abiotic columns (Fig. 5e). Copper concentrations of effluents from biotic columns were b0.03 mg/l until the pH dropped below pH 6, after which the Cu concentration plateaued just under 20 mg/l (Fig. 5f). Copper concentrations increased rapidly once the pH was below ~4.6 to give a bimodal Cu discharge pattern in the long-term column experiment (Fig. 5f). In comparison, the pH of the control column dropped steadily and once below

pH 5 a spike in Cu concentrations was observed followed by a gradual decrease in dissolved Cu (Fig. 5f). Dissolved Pb followed a similar though less pronounced trend to that of Cu (Fig. 5g). Other trace metal concentrations such as Zn, Ni and Co also show a similar trend to those of Cu and Pb, with bimodal discharge patterns for biotic columns and an earlier peak in the abiotic column (Fig. 6a–c). Trace elements such as As, Se and Cr show irregular patterns of low to very low metal release into solution (Fig. 6d–f). 4. Discussion 4.1. Bacterial population in tailings and leaching columns Our bench-top experiments confirmed that Salobo tailings, as a leaching column, can sustain large populations of A. ferrooxidans (Figs. 3; 4a–c, e), indicating that these iron oxidising bacteria are welladapted to the conditions used in this experiment. DNA sequencing of the leaching column material and MPN counts of column effluents demonstrated the abundance of A. ferrooxidans in leaching columns at pH ~5 after 185 days. Furthermore, abundant EPS and sporadic bacterial fossils hidden amongst ferrihydrite with elongated textures typical of precipitates in bacterially-mediated environments (Loiselle et al., 2018) were found in 395 day samples which reached a pH of ~3.4 (Fig. 4a, b). A section through a biotic leaching column with similar Salobo mine waste material, with higher sulphide contents and a final pH of ~2.7 after 395 days, showed pervasive A. ferrooxidans fossils within ferrihydrite precipitates (Fig. 4d), which are not visible in images of the grain surfaces in whole mounds. The abundance of fossils within this section also reflects the improved growth conditions for A. ferrooxidans at lower pH. However, A. ferrooxidans was below detection limit in the natural tailings samples, based on DNA extraction and 16S rRNA gene profiling. It is likely that A. ferrooxidans is initially sparse and only present in micro niches that were missed in the 1 g subsample for DNA extraction. Hence, given the variability between the two tailings microbial communities observed in this study (e.g., compare Tailings Samples A and B; Fig. 3) it is possible that samples that might have been rich in A. ferrooxidans were missed during sampling and to characterise the tailings community fully, more sampling and sequencing would be required. However, A. ferrooxidans is easily recovered by selective cultivation from fractures at the mine site and the sulphur oxidiser Thiobacillus thioparus is abundant (~15%) within one of the tailings samples (Fig. 3). Thiobacillus and other satellite species can have a great effect on circumneutral pH environments by changing their geochemical surroundings and establishing favourable growth conditions for other species such as A. ferrooxidans (Southam and Beveridge, 1992; Korehi et al., 2014), so more abundant colonisation of the tailings is inevitable. 4.2. Bacterial nutrients A. ferrooxidans populations gain their metabolic energy from oxidation of reduced S or Fe (Rohwerder et al., 2003). Their preferred environment for optimum activity is acidic and commonly involves oxidation of sulphide minerals (pH b 3; Amaro et al., 1991). Several studies (Bigham et al., 2001; Santelli et al., 2001; Welch and Banfield, 2002; Dopson et al., 2009; Bhatti et al., 2011a; Bhatti et al., 2011b; Henne et al., 2018) have shown that soluble Fe (II) from silicates can be used by A. ferrooxidans to meet their energy needs, and ferrous Fe is plentiful within Fe (II)-bearing silicates and magnetite in the tailings material at Salobo (Fig. 2c). Furthermore, A. ferrooxidans can populate tailings at circumneutral pH by creating microenvironments in which a low pH can be maintained (Southam and Beveridge, 1992; Mielke et al., 2003; Pace et al., 2006; Dockrey et al., 2014). It is therefore not surprising that the endemic strain that we have cultured from Salobo (Henne et al., 2018) is apparently well adapted to this Fe-rich silicate-

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(a) 0.25

(d)

0.16

*As [mg/L]

Zn [ppm]

0.2 0.15 0.1

99

0.14

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0.12

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0.1 0.08 0.06 0.04

0.05

0.02 0

0 0

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0.25 0.2 0.15 0.1 0.05 0 0

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0.5 0.45 0.4 0.35 0.3 0.25 0.2 0.15 0.1 0.05 0

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0.09 0.08 0.07 0.06 0.05 0.04 0.03 0.02 0.01 0 0 0.003 0.0025

***Cr [mg/L]

**Co [mg/L]

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0.002 0.0015 0.001 0.0005 0

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Fig. 6. Soluble elements in mg/l in leach solutions with and without A. ferrooxidans. (a) Zinc. (b) Nickel. (c) Cobalt. (d) Arsenic. (e) Selenium. (f) Chromium. * guideline values for As (V); DVG for As(III) is 24 μg/l; ** no guideline values available; *** guideline values for Cr (VI); DVG for Cr(III) is 3.3 μg/l with unknown species protection levels; DGV: Water quality toxicant Default Guideline Values in freshwater that protect 95% of species according to the Australian and New Zealand Guidelines for Fresh and Marine Water Quality (http://www.waterquality. gov.au/anz-guidelines).

dominated circumneutral environment provided by the tailings derived from this mine. In addition to energy requirements from mineral oxidation reactions, A. ferrooxidans require a range of nutrients such as P, N and C. For our experiments, these were provided in abundance in the nutrient solution, to enhance the bacterial activity in the short time frame of the laboratory environment. However, despite the regular addition of dissolved P to the columns, most of the added P was found to have deposited as a precipitate within the columns during the experiments (Figs. 4ab; 7a). The mineralogical nature of this P-rich precipitate is not known, because of its fine particle size (Fig. 7a). As a result of this P precipitation, the effluents from biotic and abiotic columns contained negligible P (Table 2; Fig. 7b). Geochemical modelling using typical effluent concentrations shows that Fe-phosphate and Ca-phosphate minerals are relatively insoluble under our experimental conditions (Fig. 7d). Dissolved Ca is derived from dissolution of Ca-bearing carbonates and other Ca-bearing minerals in the tailings (Figs. 2c; 7cd). Hence, inorganic phosphate precipitation as ferric phosphate and/or Cabearing phosphate, are apparently responsible for the depletion of dissolved P from the effluents (Fig. 7d). Furthermore, phosphate readily adsorbs to ferrihydrite (Figs. 4a; 7ae). Nevertheless, the long-term

survival of bacteria in the biotic columns indicates that adequate dissolved P was available within this geochemical system or could be readily solubilised from P-bearing minerals; A. ferrooxidans cells were found outside and within P-bearing ferrihydrite (Fig. 4b). In the field environment of the Salobo tailings impoundment, no artificial addition of nutrients will occur, so survival of these endemic bacteria requires endemic nutrients. Dissolved P contents are very low in water samples from the Salobo area (Table 1). However, P is present in the accessory mineral apatite, which in turn is present in mine rocks and tailings (Fig. 2; deMelo et al., 2016; Henne et al., 2018). Dissolution of apatite is a likely source of endemic P during long-term storage of the tailings (Fig. 7d). Low levels of N (2.64 mg/l) were found in water samples from the tailings impoundment (Table 1) and are likely derived from minor amounts of nitrogen in phyllosilicates and potentially from remnants of blasting activity. Carbon is also an essential nutrient, that A. ferrooxidans is able to fix from CO2 in the atmosphere (Kelly and Harrison, 1989) and dissolved C is available from dissolution of carbonate minerals in the tailings (Fig. 7d; Table 1). In addition, water in the Salobo tailings impoundment contains N4 mg/l dissolved organic C (Table 1), which is below the inhibitory threshold for A. ferrooxidans (Tuttle and Dugan, 1976).

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Fig. 7. Mineralogy related to phosphate in Salobo tailings. (a) Ferrihydrite precipitate in column leachate. (b) Low dissolved phosphate in column leachates despite regular addition of nutrient media. (c) Abundant dissolved Ca in leachates. (d) Modelled mineral solubility diagram (from Geochemist's Workbench; Bethke, 1998) with variable pH and dissolved P, for iron (red lines and lettering), and calcium‑magnesium (black lines and lettering, with coloured fields). Arrow indicates probable water evolution pathway within tailings. (e) Variation with pH of adsorption of dissolved P to ferrihydrite (summarised from Mengistu et al. (2015)).

4.3. Neutralisation capacity of Fe-silicates in A. ferrooxidans bearing environments The Salobo tailings material contains only minor amounts of Febearing sulphides and most of these sulphides are encapsulated in silicates (Fig. 2d–e). Such sulphides are the preferred energy sources of A. ferrooxidans (Edwards et al., 1999). Previous work by Henne et al. (2018) investigated how the endemic, iron oxidising bacteria used in this study interact with rocks from Salobo waste stock piles. The study demonstrated that A. ferrooxidans can meet their energy needs by metabolising soluble Fe (II) released from Fe (II)-bearing silicates and magnetite. The biologically-mediated weathering of these silicates buffered the leaching systems at pH ~5. This was not the case for abiotic controls nor for inoculated columns that were dominated by Fe-sulphides (Henne et al., 2018). In the present study, the increase in soluble Mn in all effluent solutions (Fig. 5c) gave the earliest indications of silicate alteration reactions in the tailings. This Mn was probably liberated during early dissolution of biotite and chlorite by both abiotic and biologically catalysed leaching (Henne et al., 2018). The concentration of dissolved Al initially lagged behind that of Mn (Fig. 5bc), due to the circumneutral pH (N5), which caused kaolinite clay alteration of the silicate minerals and/or Al-hydroxide precipitation rather than Al dissolution (Henne et al., 2018). In the present study, the pH of biotic and abiotic effluent solutions remained high initially, because of dissolution of carbonate minerals (Figs. 5a, 9d). However, after 3 months the effluent pH trends of abiotic and biotic columns diverged, with the abiotic leachates yielding progressively lower pH. In contrast, A. ferrooxidans in the biotic columns maintained the tailings leach solution pH near 5 as a by-product of silicate decomposition reactions associated with bacterial oxidation of Fe (II) to Fe (III) (Fig. 5a), similar to observations from our earlier study (Henne et al., 2018). However, the neutralisation capacity of the tailings, on the time scales of our experiments, was eventually overcome after

about a year by our on-going addition of acidic nutrient solution media, and the effluent pH decreased towards that of the media (Fig. 5a). Hence, our columns were essentially long-term acidneutralisation titration experiments, contrasting the relative effects of biotic and abiotic environments with increasing external acid addition. Our results show that without external acid addition, the long-term pH of the tailings in the tailings impoundment will remain circumneutral as water compositions are controlled by silicate dissolution reactions, and these neutralising reactions will be enhanced by the activity of A. ferrooxidans. 4.4. Storage and mobility of Fe Our medium (185 day) and long-term (395 day) leaching trials can be used to predict the rate of chemical versus biologically catalysed oxidation rates of Fe within the tailings (Fig. 8a–c). Microbial oxidation activity was indicated by low (b0.1 mg/l) dissolved Fe in effluents from biotic columns (Fig. 5d), as soluble Fe (II) is immediately oxidised to Fe (III) and precipitated as insoluble ferrihydrite (Fig. 8a–c). This was confirmed by the presence of pervasive ferrihydrite on grains in biotic columns (Fig. 4ab). The abiotic effluents initially had very low dissolved Fe (II), but rose rapidly after about three months when the pH dropped below 5 (Figs. 5a, d; 8ab). The initially low abiotic dissolved Fe prevailed because abiotic oxidation of Fe (II) is very rapid (time scale of seconds to hours) at circumneutral pH (Fig. 8b; Langmuir, 1997). However, as the abiotic pH dropped below 5, the time required for abiotic oxidation of Fe (II) increased to weeks or years (Fig. 8b). This slow oxidation rate is beyond the time scale of our experimental sampling of the effluents, so detectable Fe (II) persisted in our abiotic samples (Fig. 8a–c). Hence, our experiments clearly demonstrate the contrasts between biotic and abiotic oxidation rates for dissolved Fe (II) released during decomposition of Fe-silicates (Fig. 8a–c). Importantly, A. ferrooxidans in the biotic

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trace metals associated with sulphides, such as Pb (Fig. 5g), Zn, Ni and Co (Fig. 6a–c) attests to some sulphide alteration processes. There was negligible difference in Cu dissolution between abiotic and biotic experiments in the six-month experiments (Fig. 5f). In the longer-term biotic experiment, when the rate at which acid was added and produced overcame the effect of Fe (II)-silicates weathering, i.e., bacteria-mediated neutralisation capacity, Cu leaching was enhanced by this acidification to give a second pulse of Cu extraction (Fig. 5f). However, that enhanced Cu extraction was probably the result, in part, of inorganic reactions driven by the added acid rather than biologically-induced processes. 4.6. Storage and mobilisation of metals in tailings

Fig. 8. Summary of iron solubility and precipitation processes during tailings leaching experiments. (a) Comparison of biotic (395 days) and abiotic (185 days) iron dissolution versus pH. (b) Variations of rates of oxidation of dissolved ferrous iron with pH (after Langmuir, 1997) for biotic and abiotic systems. (c) Model iron Eh-pH diagram (Geochemist's Workbench; Bethke, 1998) showing the different Fe oxidation reactions for differing rate laws in (b).

columns accelerated the oxidation of Fe (II) and effectively immobilised the Fe within the tailings under all pH conditions (Fig. 8a–c). 4.5. Extraction of additional Cu from tailings Our medium (185 day) and long-term (395 day) leaching trials indicate some of the likely biogeochemical processes affecting the mobility of Cu into tailings waters during long-term storage, and the potential extraction of remnant Cu by bioleaching. Cu solubility is strongly dependent on pH in an oxidising environment, and is most soluble at pH b 5. Dissolved Cu content of ambient tailings water at pH N 7 is low (0.05 mg/l), and pre-mining surface waters contained b0.003 mg/l at a pH of 7 ± 1 (Konhauser et al., 1994; Damous et al., 2002; Table 1), though concentrations were slightly higher in the vicinity of the ore deposit (0.007–0.02 mg/l), which was attributed to anthropogenic activity in the area (Damous et al., 2002). In the current leaching study, Cu dissolution (Fig. 5f) was facilitated by the initial silicate alteration in biotic and abiotic experiments, as small amounts of Cu-sulphides exposed on particle surfaces, and those liberated from altered silicates, were dissolved. This sulphide dissolution is supported by data of weakly elevated levels of dissolved S in biotic and abiotic leach solutions (Fig. 5e). Likewise, minor leaching of

The tailings facility at Salobo can currently hold up to ~110 Mm3 of tailings, stretching over an area of ~6 km2 (Burns et al., 2017; Fig. 1cd). It was built in Mirim creek which drains into the Salobo River which in turn drains into the Itacaiúnas River (Fig. 1c). Natural background data from before mining production began in 2012 (Burns et al., 2017), indicate circumneutral pH (~6.8) within the Salobo Creek (Damous et al., 2002; Table 1). Waters were dominated by Ca, Na, Mg and K with low concentrations of Cu and other trace metals (Table 1; Konhauser et al., 1994; Damous et al., 2002). Concentrations did not differ significantly between dry and wet seasons (Damous et al., 2002). Our water analyses indicated slightly raised pH values compared to pre-mining values and paste pH values of tailings material indicated even higher pH (up to 8.2). This is likely due to the addition of lime during ore processing, which temporarily raises the pH to ~10 before NaHS is added to clean bornite surfaces (Vos et al., 2015; Burns et al., 2017). This is also reflected in the elevated Ca concentrations in tailings water (~15 mg/l vs. 2.1–6.5 mg/l; Table 1), which is recycled back into the processing plant (Burns et al., 2017). Our medium (185 day) and long-term (395 day) leaching trials indicate some of the likely biogeochemical processes impacting on the storage and mobilisation of metals from the tailings. As expected from the low Cu and other metal concentrations in analysed tailings material and the high pH (7.0–7.4) of water measured in the tailings dam, little soluble Cu (b0.03 mg/l) was present at circumneutral pH (Fig. 5a, f). Similarly, the levels of potentially toxic metals are below water quality toxicant default guideline values (Fig. 6; http://www.waterquality.gov. au/anz-guidelines; values are Australian/New Zealand guideline values, due to the lack of South American standard guideline value availability) and within the range of previously analysed background values (Table 1; Damous et al., 2002; Konhauser et al., 1994). Our bench-top scale experiments indicate that A. ferrooxidans can have a direct influence on the mobility - or lack thereof - of Cu and metals such as Zn, Pb and Ni within tailings. The same bacteriallymediated mechanism that hinders the bioleaching of Cu from the tailings material, by neutralising the pH, prevents acidification and limits the release of potentially toxic metals (Fig. 9). In these Fe-rich tailings, bacterial oxidation of Fe is the principal geochemical process, and this produces abundant ferrihydrite. At the circumneutral pH range expected to prevail in the silicate-rich Salobo tailings dominated by A. ferrooxidans, most trace metals are readily adsorbed on to that abundant ferrihydrite (Fig. 9; Langmuir, 1997; Drever, 1997) virtually immobilising them. Furthermore, Moon and Peacock (2013) showed that over this pH range, adsorption of trace metals, such as Cu (II), to biological ferrihydrite are higher compared to adsorption rates to chemically produced ferrihydrite. 5. Conclusions Our leaching experiments of tailings material from the Salobo mine in Brazil found that the iron oxidiser A. ferrooxidans is endemic and welladjusted to the Salobo mine environment, and tailings material in leaching columns can sustain large populations of these acidophilic iron oxidisers despite initially circumneutral pH conditions. The

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Fig. 9. Summary of the principal biogeochemical processes that will affect the tailings, contrasting the distinction between bacterially-mediated processes (centre zone) and the abiotic processes that arise when bacterial acid neutralisation is overcome by acid addition (right zone). Curves for adsorption of dissolved trace metals (bottom) have been adapted from Langmuir (1997) and Drever (1997) for the pH evolution of the experiments in this study.

leaching experiments demonstrate that A. ferrooxidans can drive geochemical processes by enhancing silicate dissolution reactions as well as promoting the precipitation of ferrihydrite. In the sulphide-poor Salobo tailings, Fe (II)-silicate dissolution reactions control pH, while ferrihydrite precipitation facilitates adsorption of metals. Both of these processes affect water composition. Silicate dissolution was indicated by an initial release of Mn and the precipitation of dissolved Al as Al hydroxides, and clay at circumneutral pH (Fig. 9). The neutralisation capacity of inoculated Fe (II)-silicate-bearing tailings was only overcome by continuous addition of acid (Fig. 9). Iron is virtually immobilised by

precipitation of ferrihydrite, both chemically at high pH and biotically at low pH, as Fe (II) oxidation rates and therefore the mobility of Fe is greatly affected by A. ferrooxidans activity (Figs. 8c; 9). The bacterially-mediated Fe-silicate driven pH buffering and the adsorption onto ferrihydrite limits the extraction of Cu, despite liberation of minor Cu from Cu-sulphide inclusions in Fe (II)-bearing silicates (Fig. 9). However, with decreasing pH adsorbed metals were likely dissolved and bacterial activity likely increased leading to raised levels of dissolved Cu in solution (Fig. 9). Low metal concentrations were also observed for other metals associated with sulphides, such as Pb, Ni and Co

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(Fig. 6a–c), which were greatly affected by cation adsorption onto ferrihydrite, the precipitation of which was especially promoted in the inoculated columns (Fig. 9). The conditions inhibiting the release of Cu, therefore also inhibit the release of potentially toxic elements (including As; Fig. 9). Supplementary data to this article can be found online at https://doi. org/10.1016/j.scitotenv.2019.04.448. Acknowledgements We thank Richard Webb and Robyn Chapman of the Australian Institute for Bioengineering and Nanotechnology, Ron Rasch and Kim Sewell of the Centre of Microscopy and Microanalysis, Elaine Whiteman of the Sustainable Minerals Institute, Julius Kruttschnitt Mineral Centre, Nathan Clayton from the Advanced Water Management Centre, Gang Xia, Paulo Vasconcelos and Marietjie Mostert from the School of Earth and Environmental Sciences and the mining staff from the Salobo mine for their technical support. 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