Prevention of Acid Rock Drainage formation through pyrite inhibition by silica coating

doi:10.21203/rs.3.rs-5200142/v1

Download PDF

Research Article

Prevention of Acid Rock Drainage formation through pyrite inhibition by silica coating

https://doi.org/10.21203/rs.3.rs-5200142/v1

This work is licensed under a CC BY 4.0 License

You are reading this latest preprint version

Passive treatment of acid rock drainage (ARD) has been considered a sustainable approach in the long term, with sulfide inhibition by silica being a promising alternative. In a small-scale column leaching, a total of four cells loaded with pyritic waste rock (11 wt.% S) from an operating Cu mine in Sweden were kept in a climatic chamber at a controlled temperature and humidity. The waste rock was subjected to a water-leach for 11 weeks before treatment using alkaline silicate solution was applied, without pH buffer and adjuster. One cell was left untreated, whereas the others were treated with silicate solution as a source of dissolved silica, with and without H₂O₂ pre-oxidation. The pH in silica-treated cells generated leachate with circumneutral pH until the end of the leaching cycle, whereas sulfide oxidation accelerated in the absence of treatment. Leachate quality in all Si-treated cells improved, as evidenced by the suppressed release of sulfur and other metals (e.g. Al, Fe, Cu, Co, Mn, and Ni).

Silica (SiO₂) layer developed on waste rock upon treatment with a longer contact time, which remained stable upon extended exposure to air and water up to 10 weeks after treatment. Pyrite inhibition is attributed to the formation of silica layer. Despite forming a siliceous Fe-O phase, H₂O₂ pre-oxidation resulted in indirect oxidation of sulfides and other phases. With an excess of silicate solution and at alkaline pH, pyrite surfaces are devoid of coating and metal ions were mobilized. Finally, this study suggested that treatment of pyritic waste rock using silica can attenuate ARD formation and prevent metal leaching by pyrite inhibition and maintaining a circumneutral pH environment or both.

acid rock drainage

pyrite

sulfur

waste rock

leaching

silica

coating

inhibition

The global demand for metals for economic and societal development, particularly clean technologies, has increased exponentially. In the last ten years, Swedish ore production has nearly doubled the production of the previous decade. In Sweden, extraction of base and precious metals, e.g., Cu, Zn, Pb and Ag, mainly originate from polymetallic sulfide-bearing ore deposits and may produce waste containing iron-sulfide minerals such as pyrite (FeS₂) and pyrrhotite (Fe(_1−x)S). In 2022, the annual generation of waste rock from Swedish non-ferrous mines was approximately 32.6 Mt (Liljenstolpe et al. 2023). Waste rock considered sub-economic (i.e., below the cut-off mineable grade) is transported and deposited in heaps, left under ambient conditions. Over time, waste rock exposed to oxygen and water enables sulfides to oxidize, and the weathering products can form Acid Rock Drainage (ARD) unless prevented in time. In Sweden, waste rock containing low concentrations of sulfides can be utilized for construction at the mine (on-site). Waste rock containing high concentrations of sulfides might be considered for waste valorization to produce sulfuric acid (H₂SO₄). Waste rock with medium sulfide content falls into neither category, thus requiring treatment to prevent ARD formation.

ARD is one of the most critical environmental problems that cause acidification and metal(loid) contamination of water and soil (Peppas et al. 2000), as well as impairs waterways (Akcil and Koldas 2006). Furthermore, the discharge of ARD into receiving water bodies causes increased turbidity and sedimentation, acid-base imbalance (Acharya and Kharel 2020) and biotic impairment, which often occurs via immediate toxicity, habitat destruction by metal precipitates, or alterations of nutrient cycle. In turn, water impairment due to ARD remains unsuitable for domestic consumption, as well as agricultural and industrial uses (Skousen et al. 2017). ARD is primarily characterized by low pH, high total dissolved solids and dissolved concentrations of metal(loid)s. As the sulfides react with oxidants (O₂ and Fe³⁺) and water, elements that are associated and hosted within them will also be released simultaneously and further render the discharge high in dissolved metal(loid)s, such as As, Co, Ni, Pb, and Zn. The acidic environment may dissolve silicate minerals and mobilize additional elements, such as Al. If the net neutralizing capacity in the waste rocks is low, presumably due to the lack of readily soluble pH buffering minerals (i.e., calcite/CaCO₃), a measure should be taken to treat the water or prevent the formation of ARD, as ARD is hard to suppress once it initiates.

During the mining operations, the most common method to limit ARD is active water treatment, i.e. treating by active addition of lime (CaO), hydrated lime (Ca(OH)₂) anhydrous ammonia (NH₃), or NaOH to neutralize acidity (Skousen et al. 1998). However, this method is not cost- and resource-efficient as it requires continuous consumption of chemical reagents, ongoing expense for operation and maintenance (Skousen et al. 2017), and considerable manpower for continuous input for alkalinity to neutralize ARD (Tu et al. 2022). In the long term, active treatment is also associated with increased maintenance and management cost (Chen et al. 2021), costly investment for membrane separation process (Aguiar et al. 2016) and handling of sludge (Tu et al. 2022) as secondary pollutant, thus is not sustainable. In water treatment, where ARD has already developed, the method aims to mitigate the effect of spreading contaminated water. Accordingly, the current best practice is at-source prevention of ARD rather than treating the water. ARD prevention methods include, for example, physical barriers, inhibition of reactive surfaces by bacteria, and chemical passivation, among others (Sahoo et al. 2013). Physical barriers, such as mine waste cover (dry or wet) over the waste (Hallberg et al. 2005; Jia et al. 2015), effectively retards sulfide oxidation and is typically applied as a measure at mine closure. However, it is a site-specific method which relies on on-site climate, hydrology, and reactivity of the mine waste (Mine Environment Neutral Drainage Program (MEND) 2004). For instance, desulphurized tailings as cover to prevent ARD have been proven suitable in a humid climate (Demers et al. 2008).

Chemical passivation, alternatively known as inhibition of pyrite surfaces, effectively reduces ARD at its source (Tu et al. 2022) and is a cost-effective method to prevent pyrite oxidation (Park et al. 2019). Inhibition is analogous to the principle of oxygen barrier coverage whereby a passivating layer attaches to the mineral surface to form a dense coating to inhibit surface oxidation, surface dissolution, or surface adsorption (Li et al. 2024) and eventually block acid production (Zhang et al., 2023). Inhibition may complement covers since it has proven effective in suppressing pyrite oxidation by reducing their reactivity. Previous inhibition research involved the use of secondary raw materials (SRM) for instance, green liquid dregs (Alakangas et al. 2013), lime kiln dust (Nyström et al. 2019a), cement kiln dust (Nyström et al. 2019b), blast furnace slag (Nyström et al. 2019b), and fly ash (Alakangas et al. 2013; Nyström et al. 2019b;rez-López et al. 2007, 2009), to neutralize pH and promote the formation of hydrous ferric oxide (HFO) which can prevent pyrite from reacting with oxidants. The amount of SRM added to the waste rock is limited to 5 wt% for the treatment to become economically feasible if transport cost is factored in (Alakangas et al. 2013, 2014).

Possible risks and concerns are raised regarding inhibition types and methods as they may be attributed to specific issues, e.g., eutrophication potential in downstream water bodies by phosphate-based coatings (Evangelou 2001; Kollias et al. 2019), or the toxicity and degradability of organic coatings by microorganisms (Jiang et al. 2000), and organosilane encapsulation methods (Dong et al. 2020, 2022) which are known to be less selective, unavailable in nature and hence expensive (Ouyang et al. 2015). Recent research in pyrite inhibition includes, for instance PropS-SH/halloysite nanotube (HNT)- Benzotriazole (BTA) (PSHB) (Li et al. 2021) and PropS-SH-tannic acid coatings (Li et al. 2023). However, these methods were applied to pure pyrite samples, hence their applicability to complex, heterogeneous pyritic samples using sources available in nature as well as their geochemical implication still require further investigation.

Concerns regarding applicability, cost and environmental feasibility raise the need to develop an inert and environmentally friendly coating from readily available natural sources. Passivation coatings by silica and silicate are more environmentally friendly (Zhang et al. 2023). In addition, silica is omnipresent in geological materials, e.g. soil, sand, suspended colloidal clays and related minerals (Iler 1979) and biogenic sources, as well as in many siliceous industrial remnants, e.g., slag and tailings. Due to its environmentally benign nature, the suitability of using silica for pyrite inhibition has garnered attention. Silica application to inhibit pyrite oxidation has been investigated previously, marked by the formation of a silicate-stabilized coating on the pyrite surface (Fan et al. 2017; Zhang and Evangelou 1998), Fe-silica passivating layer on pyrite (Kang et al. 2024; Kollias et al. 2022) that is even stable at a pH 2.5-4.0 (Evangelou 1996; Kargbo and Chatterjee 2005), and silica precipitate on pyrite at neutral pH (Bessho et al. 2011).

Despite their prevalence, comprehensive studies involving silica for pyrite inhibition primarily includes a pre-oxidation step by the addition of H₂O₂ in a buffered solution (pH 4–6) for the generation of ferric ions (Fe³⁺) (Kang et al. 2024) and later to form a silicate-stabilized ferric oxyhydroxide at a circumneutral pH. Most research on silica application for pyrite inhibition has predominantly tested on pulverized monomineralic pyrite samples (Bessho et al. 2011; Dong et al. 2020, 2022; Evangelou 2001: Fan et al. 2017; Kargbo and Chatterjee 2005) or pre-treated samples to remove secondary phases (Dong et al. 2022; Kollias et al. 2022). Therefore, the stability and sustainability of passivation technology in a complex environment is unclear (Chen et al. 2021). Additionally, previous treatment has been mainly conducted by applying coating solutions with stirring (Dong et al. 2022; Wang et al. 2019), mixed with buffer solution, such as CH₃COONa buffer (Kang et al. 2024; Kollias et al. 2022) or NH₄OH (Dong et al. 2020) for pH adjustment in order to favour the precipitation of some mineral phases (Butler and Brase 2024), which eventually does not warrant ease and suitability for field application. The examination of formed coating and effectiveness of pyrite inhibition on heterogeneous materials, as well as the geochemical characteristics of the leachate also have received limited attention, which limits the use of silica coating technologies (Butler and Brase 2024). Furthermore, the effect of repeated addition and longer contact time of silica has not been well established. Thus, to fill the research gaps, this study aims to prevent ARD formation through pyrite inhibition in waste rock by using an alkaline silicate solution as the source of dissolved silica without pH buffer and adjustment. Analyses of the solid and liquid phases, comparative geochemical characteristics of the leachate of (un)treated waste rock and implication of this study will be discussed.

Waste rock characteristics and analyses

The Kristineberg Mine, owned by Boliden Mineral AB, is situated 120 km west-northwest of Skellefteå in Västerbotten county, Sweden. Since the production began in 1940 until today, a total of 33Mt of ore have been mined with an average grade of 1.2g /t Au, 38 g/t Ag, 1% Cu and 3.8% Zn (Bjänndal and Bradley 2023). The Kristineberg Mine is an operating mine, hosting a polymetallic Volcanogenic Hosted Massive Sulfides (VHMS) deposit in the locality, which encompasses Rävliden, Rävliden North, Rävlidmyran, Hornträsksviken and Kimheden mines (Bjänndal and Bradley 2023). Waste rock containing pyrite was collected from a heap in the Kristineberg area and the samples originated from the Kristineberg Mine. The rocks were relatively fresh to slightly oxidized upon collection and had been deposited in the heap for less than one year.

The rocks sampling campaign was performed preferentially on waste rock boulders with visible sulfides and did not represent all the rock types in the heap. The expected sulfur content was approximately 10 wt%. Eight pyritic waste rock of boulder size were collected; some were selected and crushed into 4–6 mm fractions for the leaching experiment. Crushing, screening, and splitting of the waste rock for the preparation of the leaching experiment were carried out by ALS Piteå, Sweden, and the resulting particle sizes ranging from 4 to 6 mm were selected in this study for the leaching test. The remaining samples were sawed into 2 cm thick rock slabs for µ -XRF (X-ray Fluorescence) analysis, whereas another sample fraction was crushed and ground into powder for mineral characterization using XRPD (X-ray powder Diffraction). Thin sections were prepared using the remaining sample for the SEM-EDS (Scanning Electron Microscopy – Electron Dispersive Spectrometer) and LA-ICP-MS (Laser Ablation-Inductively Coupled Plasma-Mass Spectrometry) analyses.

For mineralogical analysis, rock powders were analyzed using a semi-quantitative X-Ray Powder Diffraction (XRPD) (Malvern, PANanalytical Empyrean) at Luleå University of Technology. The rock powders were examined using Cu-Kα radiation at 30 kV and 20 mA conditions and a scan speed of 0.07^o /s using a 20 mm mask. The scan time was set to 20 minutes, and the scan scope was 5^o – 90^o. Phase composition was identified by analysis of the diffraction patterns. Similarly, the formed Si precipitate was analyzed under the same conditions, except a 15 mm mask was utilized. The detection limit of XRPD with Rietveld refinement is < 1 wt% per phase. Both multi-phase identification and Rietveld refinement were performed using Profex software using BGMN database. To provide elemental distribution and relative correlation between elements, µ-XRF (Bruker M4 Tornado) at Luleå University of Technology for small spots (< 20 µm spot size) was done on rock slabs. SEM analysis (Zeiss Sigma 300 VP) was conducted equipped with an EDS detector to quantify the % mass or atom of the detected elements to analyze the morphology and identify the composition and elemental distribution of the solid phase. A high-resolution Inductively Coupled Plasma – Sector Field Mass Spectrometry (ICP-SFMS) analysis was performed at ALS Scandinavia (Luleå) to measure the major and trace elements in the rock samples. To quantify the trace elements in pyrite directly on the rock samples, analysis using LA-ICP-MS was performed at Luleå University of Technology in addition to ICP-SFMS. LA-ICP-MS offers a wide range of applications for trace element quantification, including quantitative sub-ppm trace element analysis. In contrast to ICP-SFMS analysis of the whole rock, which quantifies the elemental concentration in the bulk solid samples, LA-ICP-MS targets quantifying trace elements present within pyrite. Thus, this study added LA-ICP-MS to ensure which elements can be used as suitable tracers or proxies for sulfide oxidation.

Leaching experiment

Experimental set up

Four leaching experiments on the waste rock were carried out to investigate pyrite inhibition. This study utilized four cells (Fig. 1) (500 mL polypropylene columns, (⌀ 4 cm in diameter and 10 cm high). 250 g of sulfidic waste rock with particle size between 4 to 6 mm were placed in each cell. At the base of the column, an inert cloth material was placed to prevent the loss of fine particles and hinder rock materials from clogging the outlet and tubes. The weekly liquid-to-solid ratio was kept constant at 1 mL of milli-Q water to 1 g of rocks. All cells were placed inside a climatic chamber with a temperature of 20^oC and humidity at 60%. Wet-dry cycles were involved in each leaching cycle. Before and after treatment, the waste rock was let to oxidize before flushing was conducted and leachate was collected on the last day of every cycle. A schematic view of the experimental set-up is shown in Fig. 1. Prior to the experiment, all equipment was pre-rinsed thoroughly in 5% HNO₃ (ACS grade, Fisher Scientific). A blank sample of milli-Q water that percolated through an empty cell was submitted for analysis (Table 4).

One of the four cells, Cell A, i.e., untreated waste rock, was kept as a reference and let to oxidize with weekly flushing with Milli Q water. Three remaining cells were treated with an alkaline silicate solution at leaching cycle-11 (Table 1) after weeks of leaching using milli-Q water. A diluted sodium metasilicate pentahydrate (Na₂SiO₃.5H₂O 0.05 M) solution (i.e., alkaline silicate solution) was used as the source of dissolved silica. The alkaline silicate solution was of reagent grade (≥ 95.0% purity from solid form, Sigma Aldrich). The stock solution was prepared in a wide-mouth high-density polyethylene (HDPE) bottle to prevent possible leaching of Si from glass containers. The concentration of dissolved Si in the stock solution is 1.9 g/l based on the ICP-SFMS analysis (ALS Scandinavia, Sweden). In our experiment, no buffer solution was involved to alter the solution pH as opposed to other studies (Evangelou 2001; Fan et al. 2017; Kang et al. 2024; Kollias et al. 2022).

In Cell B, a diluted 30% hydrogen peroxide (H₂O₂) (reagent grade, Merck) was added into the waste rock to accelerate the oxidation of sulfides before Si addition. H₂O₂ was let to react with sulfides for 60 minutes, and the reaction was evident by the formation of bubbles owing to the H₂O₂ dissociation in water and the formation of rust. H₂O₂ and Na₂SiO₃.5H₂O solution were added slowly to ensure that the waste rock was flooded so that all pores were filled with the solution. In Cell C, 250 mL of fresh silicate solution was added repeatedly for three weeks. In Cell D, 250 mL of silicate solution was added once, but the solution was left to react with the waste rock for three weeks before leachate was drawn. After treatment was completed, weekly leaching was continued in all cells using milli-Q water until the leaching was terminated at week 23. This corresponded to a total of 24 leaching cycles in Cells A, B, and C and 22 leaching cycles for Cell D, with the liquid/solid ratio equal to the number of leaching cycles.

Table 1

Summary of the treatment in all cells
Cell	Description
A	Reference, untreated waste rock.
B	Pyritic waste rock pre-oxidized with 100 mL 0.2 M H₂O₂ for 1 hour, followed by addition of 150 mL Na₂SiO₃.5H₂O 0.05 M. Leaching continued using milli-Q water throughout the rest of the experiment
C	Pyritic waste rock treated with 250 mL Na₂SiO₃.5H₂O 0.05 M added every week for 3 weeks. Leaching continued using milli-Q water throughout the rest of the experiments
D	Pyritic waste rock treated with 250 mL Na₂SiO₃.5H₂O 0.05 M once. The solution was let to react with the waste rock at a longer contact time (i.e., for 3 weeks) before leachate was drawn. Leaching continued using milli-Q water throughout the rest of the experiments

Leachate sampling and analyses

Leachate from all cells was sampled weekly, and the volume of the collected leachate was measured. Subsamples were collected and analyzed for solution pH, electroconductivity (EC), and redox condition (ORP) using a portable HACH multimeter (HQ2200), including pH (± 0.02), EC (± 0.5%), and ORP electrodes (0.05%). Prior to use, all electrodes were calibrated.

A subsample of the leachate was filtered through a 0.22 µm nitrocellulose (Merck Millipore) filter into polypropylene (PP) bottles, acid-washed in 5% HNO₃ prior to use and stored in darkness and cold (4^oC). Before analysis, the sample was acidified with 1 ml nitric acid (suprapur) per 100 ml at the ALS laboratory. Major and trace element concentration was analyzed using inductively coupled plasma atomic emission spectroscopy (ICP-AES) and inductively coupled plasma sector field mass spectrometry (ICP-SFMS) at SWEDAC-accredited ALS Scandinavia in Luleå, Sweden. The analysis was performed based on US EPA Method 200.7 (modified) and 200.8 (modified) or quantitative screening analysis for 70 elements. Anions were not determined, but the sulfur measured by ICP-AES is assumed to represent sulfur in the form of sulfate, confirmed by geochemical calculations.

Geochemical calculations, including aqueous species distribution and mineral saturation state, were performed using PHREEQC version 3.7.3 (Parkhurst and Appelo 1999) and the WATEQ4F.dat thermodynamic database (Ball and Nordstrom 1991). The redox input (pe) for geochemical calculation was based on the ORP measurement on the leachate and the assumption that sulfur from ICP-SFMS represents sulfate sulfur. The calculated ionic imbalance of the whole dataset is 5.84(± 15.36).

Waste rock characteristics

The waste rock used in this study is andesite. A quantitative X-ray powder diffraction (XRPD) analysis shows that the main minerals in the waste rock are albite (7.0%), biotite (0.48%), microcline (6.7%), muscovite (12.3%), pyrite (0.61%) and quartz (36.3%). Clays, which are common weathering products of rock-forming minerals, i.e., chlorite (10.9%) and smectite (25.7%) are also present (Table 3). In general, the rocks are devoid of readily soluble minerals capable of neutralizing pH, e.g., calcite (CaCO₃) or dolomite (CaMg(CO₃)₂). However, gangue silicate minerals (e.g. chlorite) and the dissolution of this reactive silicate may provide notable amounts of alkalinity and should be accounted for (Miller et al. 2010).

Analysis by ICP-SFMS and ICP-AES (ALS Scandinavia, Luleå) reveals the rock samples' chemical composition (Table 2), which consists of 11% S, 9% Al, 13% Fe, and 60% Si, among other elements. Trace elements detected are 19 ppm Co, 38 ppm Cu, 925 ppm Mn, and 8.7 ppm Ni, among others.

Micro-XRF analysis indicated that the sulfur in the waste rock is disseminated, i.e., dispersed within the rock matrix (Fig. 2, a), with size ranging from 0.67–1.3 mm and locally forms small patches. Based on the SEM analysis, pyrite is the most abundant sulfide, mainly present as euhedral, cubic crystals (Fig. 2, b). The trace elements hosted in the waste rock were examined using various methods including µ-XRF, SEM, and LA-ICP-MS. Based on µ-XRF analysis, it was observed that all Co is hosted in pyrite; Cu coexists with Zn, whereas Mn coexists with Fe. SEM analysis shows that pyrite, as the most abundant sulfide, hosts trace elements such as light REE (Ce, La), Cu, Pb, and Te (Fig. 2) among others. LA-ICP-MS analysis of the pyrite grains showed that Co content is 134–308 ppm. Pyrite also contains Ni in the range of 20 ppm, 19 ppm Se, and 27–35 ppm Ti. Te and Pb are found together with locally high concentrations of Cu, otherwise not present. This analysis also confirmed that the pyrite can contain a high local concentration of Cu, which can be associated with elevated concentrations of As and Ni and the presence of Zn, Te, and Pb. As is found as single digit ppm level and can be present locally with higher concentration together with Mn and Te. Finally, based on the LA-ICP-MS analysis and µ-XRF, Co is an excellent tracer of pyrite oxidation. The deviation between LA-ICP-MS and ICP-SFMS was due to the whole rock digestion for total chemistry analysis, therefore, diluted. Meanwhile, in LA-ICP-MS, only the concentrations of trace elements in pyrite alone are quantified.

Table 2

Waste rock chemistry based on ICP-SFMS
Major elements		Trace elements
Element	Wt.%	Element	ppm
Al	8.81	As	1.82
Ca	0.43	Co	19.02
Fe	13.20	Cu	37.62
K	3.81	Mn	924.47
Mg	2.08	Ni	8.71
Na	0.25	Pb	4.02
S	10.82	Te	4.39
Si	59.68	Zn	178.68

Silica precipitates

Following the addition of silicate solution in Cell D, white precipitates visible to the bare eye formed on top of the waste rock and developed as layers on top of the waste rock with time. No white precipitates visible with bare eyes were observed in other Si-treated cells. As the precipitate formed in Cell D, Si from the silicate solution remained in the cell more extended than in other cells. Precipitation of silica on top of the waste rock may require a longer contact time to form a layer that is relatively stable over time.

The chemical composition and morphology of all phases were studied using SEM-EDS on representative rock samples at the end of the leaching period. Additionally, elemental distribution mapping was performed in all identified phases (Fig. 3, right-hand images). In Cell A (untreated waste rock), pyrite was intensely oxidized, as shown by the corroded surface and dissolution features (Fig. 3). In all cells with Si-treated rocks, Si was observed to be associated with secondary precipitates, but with differences in the occurrence. In Si-treated waste rock pre-treated with H₂O₂ (Cell B), pyrite was partly oxidized, as marked by the presence of Fe-O phase (i.e., Fe(oxyhydr)oxide), as opposed to the findings by Kollias et al. (2018), where no iron hydroxide phases were observed as separate precipitates. Silicon was detected within the Fe-O phase by SEM, indicating that Si was absorbed into this phase (Fig. 3). The presence of Si in the amorphous iron (oxyhydr)oxide has been reported to stabilize the passivating layer on pyrite (Fan et al. 2017). Silicate ions form coatings through reaction with the OH groups of ferric hydroxides on the surface of pyrite (Park et al. 2019). It has been understood that silicate species may polymerize at the surface of ferrihydrite (5Fe₂O₃.9H₂O) (Swedlund et al. 2009; Vempati et al. 1990), whereas natural siliceous ferrihydrite may exist in deposits of mine drainage waters (Cismasu et al. 2014). Further, Lee et al. (2011) documented that forming the Fe-silicate complex reduced the activity of oxidants for pyrite, thus reducing pyrite oxidation. In Cell C, a precipitate did not form on the rocks despite adding an alkaline silicate solution for three consecutive weeks (Fig. 3).

In Cell D, subjected to single Si treatment, a homogeneous layer with a composition corresponding to the stoichiometric ratio Si:O = 1:2 (i.e., silica) precipitated on the surface of waste rock, covering the sulfides and all other phases (Fig. 4). The formed SiO₂ precipitate by passivating pyrite isolates pyrite surfaces from exposure to oxidants and water (Johnson and Hallberg 2005). This precipitate remained relatively stable following weekly flushing with Milli Q water in each cycle and even after the leaching was terminated 10 weeks after treatment in Cell D. The formation of chemically insoluble, inert, and protective surface coating on the pyrite surfaces is desired for AMD control (Fan et al. 2021). However, future studies are required to test the stability of the formed coating under exposure to actual mining environments. Silica precipitate in Cell D conforms with the finding by Kollias et al. (2018), whereby pyrite treated with higher Si concentration (i.e., > 0.1 mM Si), Si tends to form stable SiO₂ precipitates, in comparison to 0.1 mM Si where it seems to favour the formation of Fe oxyhydroxides and adsorbed silicate species (Kollias et al. 2018). The generally approved mechanism for silica precipitation is the polymerization of monosilicic acid (i.e., soluble form of silica, Si(OH)₄) to form silicate oligomers which often occurs spontaneously at concentrations exceeding 100–200 ppm, followed by subsequent dehydration to form silica (Dyer et al. 2010; Iler 1979).

Based on quantitative XRD analysis (Fig. 5), the formed precipitate comprises 75.40 wt.% of amorphous silica and 24.60 wt.% of more crystalline silica. The precipitate is composed of 46.74 wt.% Si and 53.26 wt.% O, which results in the stoichiometric ratio of Si:O = 1:2. It is possible that a longer contact time between the alkaline silicate solution and the waste rock allowed for the evaporation and higher concentration of silica, which further led to the polymerization of silica to a more stable, crystalline state as evidenced by the silica layer to remain longer on top of the waste rock longer compared to other treatment.

Leaching characteristics

The following presents and discusses the leaching characteristics of the untreated and Si-treated waste rock with respect to main field parameters (pH, EC, Fig. 6) and selected major and trace elements (Fig. 7 and Fig. 8, respectively).

During the first 11 weeks of leaching and before the treatment with silicate solution was initiated, leachate from all cells had a pH, i.e., ≈ 6. In the cell with untreated waste rock (Cell A), pH declined to 4 over time as leaching progressed (Fig. 6), and the concentration of major elements (Al, Fe, S and Si, Fig. 7) and trace elements (Co, Cu, Mn, and Ni, Fig. 8) in Cell A increased progressively. These observations indicated ongoing pyrite oxidation in the untreated waste rock. Cobalt is hosted principally in pyrite and, therefore, a good fingerprint for ongoing pyrite oxidation, confirmed by the elevated release of Co from the leachate in Cell A (Fig. 8, Fig. 11).

The concentrations of major and trace elements in the leachate peaked simultaneously with pH and EC peaks and thereafter decreased over time. In all Si-treated cells, the leachate pH and EC peaked upon adding Si solution, reaching around 10 and 11 mS/cm, respectively, consistent with the alkaline pH and the high dissolved Si and Na concentrations of Si solution. As leaching continued, the pH in all Si-treated cells decreased steadily and stabilized until the end of the leaching period, but both pH and EC remained higher until the end compared to the start of the experiment (Fig. 6). At the end of the leaching period, in all Si-treated cells, the leachate pH stabilized around 7, and the concentrations of major (Al and Fe) and trace elements in pyrite (Co, Cu, Mn, Ni) were lower compared to Cell A. The slow pH decrease and lower release of metals in all treated waste rock is attributed to neutralization released upon addition of alkaline silicate solution or surface passivation of pyrite. The pre-oxidized waste rock (Cell B) systematically showed the most prominent release of Al, Fe, and S to the leachate than Si-treated cells without H₂O₂ pre-oxidation (Cell C and Cell D) (Fig. 7). No apparent differences between the leaching characteristics from Cell C and Cell D, i.e. waste rock subjected to multiple and single Si-treatment, respectively, were observed. Leaching characteristics of the waste rock treated in different ways in cells B, C, and D showed some differences and similarities. Firstly, peaking, with the highest concentrations of elements, was systematically observed in Cell B than in Cell C and Cell D (Fig. 7 and Fig. 8). The concentrations of most major and trace elements showed a decreasing trend weeks after treatment. Cell B showed some exceptions, including S, a somewhat ambiguous trend, and higher variability than those observed in other Si-treated cells (C and D) (Fig. 7; Fig. 8; Fig. 11). Impurities of the reagents used in the treatments may have played a negligible role in the peaking concentrations (Table 4), thus it seems that the pre-treatment with H₂O₂ appears not only to result in dissolution of pyrite (Eq. 2) but also to indirect dissolution of other phases, e.g., rock-forming silicates (Eq. 3; Eq. 4), hence releasing several major and trace elements into the leachate. All Si-treated cells generated and maintained a neutral pH in the leachate, with no signs of accelerated sulfide oxidation at the end of the leaching cycle.

The likely explanation for the peaking major and trace cation concentrations upon Si addition is their solubilization due to alkaline pH, and in the case of Cell B, the oxidation of pyrite and dissolution of rock-forming minerals upon H₂O₂ treatment. Upon addition of H₂O₂, dissociation of H₂O₂ in water released O₂ (Eq. 1) and led to elevated concentration of dissolved O₂ in the solution in Cell B. Due to a higher concentration of available oxidants, on-going pyrite oxidation accelerated (Eq. 2), releasing more protons (H⁺) into the solution. Acidity (H⁺) is consumed during the weathering process of rock-forming silicates, for example, albite (NaAlSi₃O₈) (Eq. 3) and relatively reactive chlorite (Miller et al. 2010) (Eq. 4), as reflected in the release and peak concentration of Al in the leachate following H₂O₂ addition (Fig. 7).

In alkaline pH (≈ 10), Al and Fe may solubilize due to the formation of anionic species Al(OH)₄⁻ and Fe(OH)₄⁻ (e.g. Bhattacharya 2013). Geochemical calculations of aqueous speciation in the leachate upon Si-treatment confirmed that the Al chemical species in all Si treated cells (cells B, C, D) are Al(OH)₄⁻, accounting for nearly 100% of the total dissolved Al. Without any treatment, 97–98% of the total dissolved Al was present as Al(OH)₄⁻ while the remaining 2% were in the form of Al(OH)₃ and Al(OH)₂⁺. In the case of Fe, the main aqueous Fe species in all Si-treated cells was Fe(OH)₄⁻, whereas Fe(OH)₃ was the dominant aqueous Fe species in the untreated waste rock (Fig. 9). Hence, it is likely that the solubilization of Al and Fe upon the addition of silicate solution rendered the Al and Fe concentrations higher in the leachate. H₃SiO₄⁻ was the main Si species in highly alkaline pH (≈ 10–11), resulting from the dissociation of H₄SiO₄. As the pH decreased and stabilized to a circumneutral level, H₄SiO₄ became the main Si species (Fig. 9). A similar mechanism is a likely explanation for the peaking concentrations S and other cations, too. The opposing trend of S in all Si-treated cells, compared to the typical decreasing trend as leaching continued after Si treatment, may also arise from the aqueous speciation and enhanced S solubility due to treatment with alkaline Si-solution. Based on the aqueous speciation calculations, dissolved sulfur existed as sulfate (SO₄²⁻) and soluble MgSO₄ and KSO₄ species in the leachate from the untreated waste rock. In comparison, SO₄²⁻ and NaSO₄ predominated in the leachate following the addition of alkaline silicate solution, whereas nearly all sulfur was present as sulfate in the leachate of all Si-treated waste rock at the end of the leaching cycle.

The decreasing trend in the concentrations of major (Al, Fe) and trace elements (Co, Mn, Pb) hosted in pyrite and the neutral leachate pH towards the end of the leaching may be an indication of suppressed pyrite oxidation or due to a circumneutral pH environment causing metals and metalloids to be captured in the secondary phases, or even both. When pH is raised, Fe oxy(hydr)oxides may precipitate, and trace elements (e.g., Cu, Mn, Ni, Zn) may be retained in these phases (Shi et al. 2021). The saturation indices (Fig. 10) showed that leachate samples from the Si-treated cells (cells B, C, D) were supersaturated or close to saturation with respect to Fe(OH)₃ and Al(OH)₃ following the treatment of waste rock with silicate solution. These phases may have precipitated in Si-treated waste rock, hence immobilizing Fe and Al.

At the end of the leaching period, the mean percent reduction in dissolved elemental concentrations in the leachate from cells B, C, and D, compared to the reference (Cell A) is reported as 94.74% ± 3.54% (Al), 98.89% ± 0.90% (Co), 91.79%±10.55% (Cu), 77.85% ± 32.3% (Fe). In Cell D, the reduction in sulfur release in the leachate is 75.2%, compared to 39.3% and 41.6% reduction in sulfur release from Cell B and Cell C, respectively. This result is in good agreement with Lee et al (2011) that documented the lower sulfate concentration in the solution when pyritic rocks were treated with H₂O₂ and sodium silicate (Lee et al. 2011), 72% reduction in sulfate release in pyritic tailings treated with Na₂SiO₃/ H₂O₂/CH₃COONa (Kollias et al. 2021), ≈ 12% (± 22%) to ≈ 49% (± 24%) reduction in sulfate on a field-scale treatment of silicate solution on a pyritic coal spoil (Vandiviere and Evangelou 1998), 40–45% reduction in sulfidic slates (Van den Eynde et al. 2009), and ≈ 35% reduction in sulfur release from the nickel-bearing waste rock treated with Na₂SiO₃/ H₂O₂/NaHCO₃ buffer solution (Roy et al. 2020).

Leachate quality profile, coupled with geochemical calculations, was used to describe geochemical reactions occurring in untreated and silicate-treated waste rock. Saturation indices (SI) computed with WATEQ4F for chemical analysis from the leachates are analyzed whether they are at, below, or above saturation with respect to possible secondary minerals.

Figure 10 shows the range of SI values for amorphous SiO₂, amorphous Fe(OH)₃ (ferrihydrite) and selected Al-phases, including amorphous and crystalline Al(OH)₃ (gibbsite), AlOOH (boehmite), Al₂Si₂O₅(OH)₄ (basaluminite) over time. The SI confirms supersaturation of amorphous SiO₂ in all Si-treated cells upon addition of silicate solution, followed by decreasing SI close to saturation, suggesting that it likely precipitated and then controlled Si solubility. In cells B and C, the silica dissolved over time, whereas in Cell D, silica remained in the cell longer than in other Si-treated waste rock. Therefore, weeks following treatment, the solution was undersaturated with respect to amorphous silica. Ferrihydrite and gibbsite may have precipitated in all cells, and the former is a controlling phase in Cell D. Amorphous Al(OH)₃ approached saturation and may have controlled the solubility of A in Cell A, as opposed to the Si-treated waste rock, whereas boehmite (AlOOH) approached saturation in Cell A and Cell D, before and after Si-treatment for the latter. Basaluminite (Al₄(OH)₁₀SO₄), a nanocrystalline aluminumoxyhydrosulfate commonly present in areas affected by ARD (Carrero et al. 2017), may have precipitated in Cell A. As the pH became more acidic in Cell A at the end of the leaching period, precipitation of basaluminite indicated an ongoing pyrite oxidation.

Prevention of ARD formation by silica treatment in waste rock

Single Si treatment of waste rock (Cell D) with a longer contact time showed the most promising results in preventing ARD from sulfidic, acid-producing waste, as evidenced by the homogenous SiO₂ coating observed on pyrite surfaces and the decreasing trends in the release of metals. This coating layer on pyrite likely acted as a protective barrier against further oxidation and could explain the nearly constant pH at the end of the leaching cycle. The formed precipitate in Cell D as layers on top of the waste rock might be important in protecting the pyrite from oxidation, as this layer was not leached at the same rate as in other cells. Furthermore, as the precipitate formed, Si from the silicate solution remained in Cell D longer than other cells. This study suggested that silica precipitation may require prolonged time to develop on the surface and finally form a stable layer over time. In addition, adding silicate solution at a slower flow rate might also be required to promote a progressive supersaturation buildup without achieving a highly alkaline pH level, since H₄SiO₄ remains undissociated at pH values below 9 and it is a precursor to solid, amorphous silica (SiO₂) (Iler 1979).

It is also worth noting that this study showed that the effect of contact time of solution is an essential factor in the development of silica layer. The contact time of the silicate solution in Cell D was more prolonged than in Cell B and Cell C to provide sufficient time for the buildup of the SiO₂ coating which may grow thicker and denser over time. Due to a longer contact time between the alkaline silicate solution and the waste rock in Cell D, evaporation may have also resulted in higher silica concentration. Once the concentration exceeded 100 ppm, the monosilicic acid either precipitated or polymerized to form silica layer on top of the waste rock (Iler 1979). This study revealed that even in alkaline solution silica precipitation is not immediately complete but requires several days to attain a steady state.

Nonetheless, despite the SiO₂ formed, it remains uncertain how long the effect of Si-treatment may last and whether the formed precipitate is stable under a more acidic environment. The decreasing trend in the release of metals was observed after Si treatment, followed by initially elevated concentrations of some elements. However, the sulfur release did not show a similar trend, which is believed to be partially due to the formation of soluble NaSO₄⁻ species from the alkaline Si solution. The current study does not unambiguously show whether the decreasing trends in the release of metals are due to the inhibition of the pyrite surface by precipitation of secondary minerals, the maintenance of a circumneutral pH environment created by the Si treatment, or the combination of both. However, pH sustenance to a circumneutral level by silicate solution is also essential because acid pH conditions enhance the mobility of metals, particularly the divalent cations of Cu, Zn, and Mn (e.g. Dold 2017).

Repeatedly adding silicate solution to the pyritic waste rock (Cell C) did not improve the overall leachate quality compared to waste rock subjected to single Si treatment (Cell D). Further, it did not result in the formation of homogeneous precipitate on the rocks; it only resulted in partial coverage of the pyrite surface. A likely explanation is that the repeated addition of alkaline silicate solution resulted in an alkaline pH > 9 for an extended time, which in turn favors H₄SiO₄ dissociation in alkaline solution, giving up H⁺ (Iler 1979). It is understood that the solubility of amorphous silica is little affected by changes of pH in the range 2–9 but increases rapidly as the pH rises above 9 (Iler 1979). The solution pH in Cell C was maintained at ~ 10.5 (Fig. 6), where the rate of silica dissolution, which is also catalyzed by hydroxyl ions, becomes significant (Nordström et al. 2011). Repeated addition of silicate solution also resulted in a possible risk of metal mobilization, as shown in elevated concentrations and leached mass of dissolved Al and Fe (Fig. 7; Fig. 11), persisting over an extended time. Given that repeated treatment did not result in further improvement compared to a single treatment but increased the possible risk for metal leaching due to alkaline pH and additional reagents required, such an approach is not motivated under the circumstances prevailing in this study.

Although the siliceous Fe(oxyhydr)oxide phase is detected in Cell B, the pre-oxidation with H₂O₂ before Si-treatment did not improve the impact of Si-treatment compared to treatment with alkaline Si-solution only (cells C and D), but rather the opposite. The pre-oxidation step resulted in prominent release of sulfur as well as several major and trace elements (Fig. 7 and Fig. 8) upon treatment, as well as showed an ambiguous trend, rather than decreasing trend, in their release to the leachate upon continued leaching. Previous Si-inhibition studies involve H₂O₂ pre-oxidation prior to the addition of silica solution to form a coating layer on pure pyrite samples (Fan et al. 2017; Kollias et al. 2018, 2022; Vandiviere and Evangelou 1998), while the current experiments on waste rock indicate that the pre-oxidation step may not have a positive effect on the leachate quality when applied on to heterogeneous materials, e.g. pyritic waste rock. This finding agrees with the results by Kang et al. (2024), which shows that Fe-silicate-based treatment using H₂O₂ pre-oxidation did not properly inhibit sulfidic rock samples, although it appears effective in treating pulverized monomineralic pyrite samples.

Accordingly, the one-time addition of silica on sulfidic waste rock with a longer contact time (Cell D) generated near-neutral-alkaline pH with low metal release. Pre-oxidation of waste rock with H₂O₂ prior to Si treatment (Cell B) or repeated addition of silicate solution to treat the pyritic waste rock (Cell C) did not further improve the overall leachate quality or precipitate a protective layer on the pyrite surface in comparison to single Si-treatment of waste rock (Cell D). In relevance to industrial application, this study serves as a precursor to further experimental work on promoting the formation of Si coatings but with silica-bearing industrial remnants, e.g. slag, or enhanced in-situ dissolution of reactive silicate-bearing minerals, e.g. as a potential source of dissolved Si. In the actual mine environment, micas and chlorite, with silica and silicate being the main components, remain stable under mine conditions and maintain the pH (Zhang et al. 2023). Silica is insoluble in acid (pH > 2). It is, therefore, unaffected by the acidification of the mine environment to form a coating on pyrite, despite requiring confirmation to examine the formation and stability of the silica precipitate under long term neutral pH, which further studies in the future must confirm. Finally, this study's outcome contributes to understanding the geochemical implication of pyrite inhibition by silica, particularly in freshly dumped pyritic waste rock at a waste heap pile in an operating mine site. To warrant applicability in the mine environment, the amount of material required for achieving inhibition of pyrite oxidation prior to field applications must be evaluated in further studies.

In a small-scale column leaching study, the treatment of sulfidic waste rock with an alkaline silicate solution managed to generate leachate with a lower concentration of metals while maintaining a circumneutral pH, as opposed to untreated rocks. Without any treatment, sulfide oxidation accelerated, resulting in low pH and elevated concentration of metals. A single addition of alkaline silicate solution with a longer contact time, without H₂O₂ pre-oxidation, resulted in homogenous precipitation of silica (SiO₂) layer on the rocks. The formed silica precipitate remained stable upon weekly dry-wet cycle until the end of the leaching period. On the other hand, H₂O₂ pre-oxidation prior to silicate addition resulted in the formation of a Si-adsorbed Fe(oxyhydr)oxide phase but also rendered the concentrations of metals high in the leachate, presumably indicating the indirect oxidation of pyrite and weathering of rock-forming silicate minerals through acid-consuming reaction. Although previous research suggested the pre-oxidation with H₂O₂ prior to silicate treatment on monomineralic pyrite samples to release ferric (Fe³⁺) ions, this does not warrant suitability when applied to complex and heterogeneous samples, such as pyritic waste rock, as it indirectly dissolves partly not only pyrite but also other phases. Our study suggested that a single addition of alkaline silicate solution with a longer contact time with the waste rock resulted in the development of SiO₂ precipitate, which remained stable upon extended exposure to air and water up to 10 weeks after treatment. The longer contact time allowed silica layer to develop over time. In contrast, repeated addition of silicate solution resulted in ongoing dissolution of SiO₂ at a pH above 10 and led to possible risk of metal mobilization in the solution. This study suggested the role of silica in preventing ARD formation, by maintaining a circumneutral pH environment over an extended time and inhibiting the surface of pyrite by precipitation of a secondary mineral in the form of SiO₂, or a combination of both. Finally, this study serves as a precursor to future experiments to inhibit pyrite surfaces using siliceous industrial by-products or enhanced in-situ dissolution of reactive silicate minerals.

Ethical approval

Not applicable.

Consent to participate

All authors have agreed to contribute in this study.

Consent for publication

The authors agree to publish.

Competing interests

The authors declare no competing interests that are relevant to the content of this article.

Funding

This study was supported and funded by Boliden Mineral AB.

Author Contribution

All authors contributed to the study conception and design. Dantie Claudia Butar Butar is responsible for material preparation, data curation and acquisition, analysis, and writing – original draft. Lena Alakangas is responsible for methodology, analysis, interpretation, review, and editing. Hanna Kaasalainen is responsible for methodology, data analysis and interpretation, review, and editing. Erik Ronne is responsible for review, editing, funding acquisition and validation. All authors commented on previous versions of the manuscript. Finally, all authors read and approved the final manuscript.

Acknowledgements

Boliden Mineral AB is gratefully acknowledged for fully funding this research, providing access during site visits, and supplying rock samples as well as other research materials. Dr. Anton Andersson is gratefully thanked for his assistance with LA-ICP-MS.

Data availability

The data that support the findings of this study are available within the paper. Should any raw data files be needed in another format they are available upon request to the corresponding author.

Acharya, B. S., & Kharel, G. (2020). Acid mine drainage from coal mining in the United States – An overview. Journal of Hydrology 588, 125061. https://doi.org/10.1016/j.jhydrol.2020.125061.
Aguiar, A. O., Andrade, L. H., Ricci, B. C., Pires, W. L., Miranda, G. A., & Amaral, M. C. (2016). Gold acid mine drainage treatment by membrane separation processes: An evaluation of the main operational conditions . Separation and Purification Technology 170, 360-369. https://doi.org/10.1016/j.seppur.2016.07.003.
Akcil, A., & Koldas, S. (2006). Acid Mine Drainage (AMD): causes, treatment and case studies. Journal of Cleaner Production, 14(12–13), 1139-1145, ISSN 0959-6526, https://doi.org/10.1016/j.jclepro.2004.09.006.
Alakangas, L., Andersson, E., & Mueller, S. (2013). Neutralization/prevention of acid rock drainage using mixtures of alkaline by-products and sulfidic mine wastes. Environmental Science and Pollution Research 20, 7907–7916. https://doi.org/10.1007/s11356-013-1838-z.
Alakangas, L., Maurice, C., Mcsik, J., Nyström, E., Sandström, N., Andersson-Wikström, A., & Hällström, L. (2014). Kartläggning Av Restprodukter För Efterbehandling Och Inhibering Av Gruvavfall: Funktion Tillgång Och Logistik. Luleå: Luleå University of Technology. Retrieved from https://www.diva-portal.org/smash/get/diva2:997612/FULLTEXT01.pdf , accessed on 24 April 2024, in Swedish.
Ball, J., & Nordstrom, D. (1991). User’s manual for WATEQ4F with revised thermodynamic database and test cases for calculating speciation of major, trace and redox elements in natural waters, pp. 91-183. Denver: U.S. Geological Survey Open-File Report.
Bessho, M., Wajima, T., & Ida, T. (2011). Experimental study on prevention of acid mine drainage by silica coating of pyrite waste rocks with amorphous silica solution. Environmental Earth Science, 64, 311-318. https://doi.org/10.1007/s12665-010-0848-0.
Bhattacharya, P. (2013). Pilot scale evaluation of soil washing for treatment of arsenic contaminated soil (PSEMA) Final Report. Formas Research Project Dnr: 245-2006-964.
Bjänndal, E., & Bradley, J. (2023, December 31). Mineral Resources and Mineral Reserves. Retrieved from Boliden: https://www.boliden.com/operations/exploration/mineral-reserves-and-mineral-resources/
Butler, B., & Brase, L. (2024). Critical Review of Field Studies of Chemical Surface Coatings to Mitigate Leaching from Mining Wastes. Mine Water Environment 43, 3-15. https://doi.org/10.1007/s10230-024-00973-7.
Carrero, S., Fernandez-Martinez, A., Pérez-López, R., & Nieto, J. M. (2017). Basaluminite Structure and its Environmental Implications. Procedia Earth and Planetary Science 17, 237-240. https://doi.org/10.1016/j.proeps.2016.12.080.
Chen, G., Ye, Y., Yao, N., Hu, N., Zhang, J., & Huang, Y. (2021). A critical review of prevention, treatment, reuse, and resource recovery from acid mine drainage. Journal of Cleaner Production 329, 129666. https://doi.org/10.1016/j.jclepro.2021.129666.
Cismasu, A. C., Michel, F. M., Tcaciuc, A. P., & Brown, G. E. (2014). Properties of Impurity-Bearing Ferrihydrite III. Effects of Si on the Structure of 2-Line Ferrihydrite. Geochimica et Cosmochimica Acta 133, 168−185. https://doi.org/10.1016/j.gca.2014.02.018.
Demers, I., Bussière, B., Benzaazoua, M., Mbonimpa, M., & Blier, A. (2008). Column test investigation on the performance of monolayer covers made of desulphurized tailings to prevent acid mine drainage. Minerals Engineering 21(4), 317-329. https://doi.org/10.1016/j.mineng.2007.11.006.
Dold, B. (2017). Acid rock drainage prediction: A critical review. Journal of Geochemical Exploration 172, 120-132. https://doi.org/10.1016/j.gexplo.2016.09.014.
Dong, Y., Liu, Z., Liu, W., & Lin, H. (2022). A new organosilane passivation agent prepared at ambient temperatures to inhibit pyrite oxidation for acid mine drainage control. Journal of Environmental Management 320, 115835. https://doi.org/10.1016/j.jenvman.2022.115835.
Dong, Y., Zeng, W., Lin, H., & He, Y. (2020). Preparation of a novel water-soluble organosilane coating and its performance for inhibition of pyrite oxidation to control acid mine drainage at the source. Applied Surface Science 531, 147328. https://doi.org/10.1016/j.apsusc.2020.147328.
Dyer, L. G., Fawell, P. D., Newman, O. M., & Richmond, W. (2010). Synthesis and Characterisation of Ferrihydrite/Silica Co-Precipitates. Journal of Colloid and Interface Science 348, 65-70. https://doi.org/10.1016/j.jcis.2010.03.056.
Evangelou, V. (1996). Oxidation proof silicate surface coating on iron sulfides. US Patent No. 5, 494, 703.
Evangelou, V. (2001). Pyrite microencapsulation technologies: Principles and potential field application. Ecological Engineering, 17(2–3), 165-178. https://doi.org/10.1016/S0925-8574(00)00156-7.
Fan, R., Qian, G., Short, M. D., Schumann, R. C., Brienne, S., Smart, R. S., & Gerson, A. R. (2021). Passivation of pyrite for reduced rates of acid and metalliferous drainage using readily available mineralogic and organic carbon resources: A laboratory mine waste study. Chemosphere 285, 131330. https://doi.org/10.1016/j.chemosphere.2021.131330.
Fan, R., Short, M. D., Zeng, S.-J., Qian, G., Li, J., Schumann, R. C., . . . Gerson, A. R. (2017). The Formation of Silicate-Stabilized Passivating Layers on Pyrite for Reduced Acid Rock Drainage. Environmental Science & Technology, 51(19), 11317-11325. https://doi.org/10.1021/acs.est.7b03232.
Hallberg, R., Granhagen, J., & Liljemark, A. (2005). A fly ash/biosludge dry cover for the mitigation of AMD at the Falun mine. Geochemistry 65, 43-63. https://doi.org/10.1016/j.chemer.2005.06.008.
Iler, R. (1979). The Chemistry of Silica: Solubility, Polymerisation Colloid and Surface Properties and Biochemistry. New York: John Wiley & Sons.
Jia, Y., Maurice, C., & Öhlander, B. (2015). Metal mobilization in tailings covered with alkaline residue products: results from a leaching test using fly ash, green liquor dregs, and lime mud. Mine Water Environment 34(3), 270-287. https://doi.org/10.1007/s10230-014-0317-1.
Jiang, C., Wang, X., & Parekh, B. (2000). Effect of sodium oleate on inhibiting pyrite oxidation. International Journal of Mineral Processing, 58(1–4), 305-318, ISSN 0301-7516.
Johnson, D. B., & Hallberg, K. B. (2005). Acid mine drainage remediation options: a review. Science of The Total Environment 338(1-2), 3-14. https://doi.org/10.1016/j.scitotenv.2004.09.002.
Kang, C., Kang, J., & Kim, K. (2024). Inhibition of acid rock drainage with iron-silicate or phosphate film: in rainy and submerged environments. . Environ Geochem Health 46, 216, https://doi.org/10.1007/s10653-024-01996-3.
Kargbo, D., & Chatterjee, S. (2005). Stability of silicate coatings on pyrite surfaces in a low pH environment. Journal of Environmental Engineering 131(9), 1340-1349. https://doi.org/10.1061/(ASCE)0733-9372(2005)131:9(1340).
Kollias, K., Mylona, E., & Papassiopi, N. (2018). Environmental Science and Pollution Research 25. Development of silica protective layer on pyrite surface: a column study, 26780–26792. https://doi.org/10.1007/s11356-017-0083-2.
Kollias, K., Mylona, E., & Papassiopi, N. (2022). Application of Silicate-Based Coating on Pyrite and Arsenopyrite to Inhibit Acid Mine Drainage. . Bull Environ Contam Toxicol 108, 532–540. https://doi.org/10.1007/s00128-021-03310-8.
Kollias, K., Mylona, E., Adam, K., Chrysochoou, M., Papassiopi, N., & Xenidis, A. (2019). Characterization of phosphate coating formed on pyrite surface to prevent oxidation. Applied Geochemistry 110, 104435. https://doi.org/10.1016/j.apgeochem.2019.104435.
Lee, J.-S., Chon, C.-M., & Kim, J.-G. (2011). Suppression of Pyrite Oxidation by Formation of Iron Hydroxide and Fe(III)-silicate Complex under Highly Oxidizing Condition. Korean Journal of Soil Science and Fertilizer, 297-302. https://doi.org/10.7745/kjssf.2011.44.2.297.
Li, D., Chen, X., Liu, C., Tian, J., Li, F., & Liu, Y. (2023). Suppression of pyrite oxidation by co-depositing bio-inspired PropS-SH-tannic acid coatings for the source control acid mine drainage. Science of The Total Environment 862, 160857. https://doi.org/10.1016/j.scitotenv.2022.160857.
Li, D., Gong, B., Liu, Y., & Dang, Z. (2021). Self-healing coatings based on PropS-SH and pH-responsive HNT-BTA nanoparticles for inhibition of pyrite oxidation to control acid mine drainage. Chemical Engineering Journal 415, 128993. https://doi.org/10.1016/j.cej.2021.128993.
Li, Y., Cao, Y., Ruan, M., Li, R., Bian, Q., & Hu, Z. (2024). Mechanism and In Situ Prevention of Oxidation in Coal Gangue Piles: A Review Aiming to Reduce Acid Pollution. Sustainability 16, 7208. https://doi.org/10.3390/su16167208.
Liljenstolpe, C., Hamberg, R., Larsson, D., & Abrahamsson, S. (2023). Statistics of the Swedish Mining Industry 2022. ISSN 0283-2038. Geological Survey of Sweden (SGU).
Miller, S., Stewart, W., Rusdinar, Y., Schumann, R., Ciccarelli, J., & Li, J. (2010). Methods for estimation of long-term non carbonate neutralization of acid rock drainage. Science of the Total Environment 408, 2129–2135. https://doi.org/10.1016/j.scitotenv.2010.01.011.
Mine Environment Neutral Drainage Program (MEND). (2004). Design, construction and performance monitoring of cover systems for waste rock and tailings, MEND Report No 2.21.4. Vancouver: Natural Resources Canada.
Nordström, J., Nilsson, E., Jarvol, P., Nayeri, M., Palmqvist, A., Bergenholtz, J., & Matic, A. (2011). Concentration- and pH-dependence of highly alkaline sodium silicate solutions. Journal of Colloid and Interface Science, 37-45. https://doi.org/10.1016/j.jcis.2010.12.085.
Nyström, E., Kaasalainen, H., & Alakangas, L. (2019). Prevention of sulfide oxidation in waste rock by the addition of lime kiln dust. Environmental Science and Pollution Research, 26, 25945–25957. https://doi.org/10.1007/s11356-019-05846-z.
Nyström, E., Kaasalainen, H., & Alakangas, L. (2019). Suitability study of secondary raw materials for prevention of acid rock drainage generation from waste rock. Journal of Cleaner Production, 575-586. https://doi.org/10.1016/j.jclepro.2019.05.130.
Ouyang, Y., Liu, Y., Zhu, R., Ge, F., Xu, T., Luo, Z., & Liang, L. (2015). Pyrite oxidation inhibition by organosilane coatings for acid mine drainage control. Minerals Engineering 72, 57-64. https://doi.org/10.1016/j.mineng.2014.12.020.
Park, I., Tabelin, C. B., Jeon, S., Li, X., Seno, K., Ito, M., & Hiroyoshi, N. (2019). A review of recent strategies for acid mine drainage prevention and mine tailings recycling. Chemosphere 219, 588–606. https://doi.org/10.1016/j.chemosphere.2018.11.053.
Parkhurst, D., & Appelo, C. (1999). User’s guide to PHREEQC (version 3) – a computer program for speciation, batch-reaction, one-dimensional transport, and inverse geochemical calculations, pp. 99-4259. Denver, Colorado: U.S. Geological Survey Water-Resour.Invest. Report.
Peppas, A., Komnitsas, K., & Halikia, I. (2000). Use of organic covers for acid mine drainage control. Minerals Engineering, 13(5), 563-574. https://doi.org/10.1016/S0892-6875(00)00036-4.
Pérez-López, R., Cama, J., Nieto, J. M., & Ayora, C. (2007). The iron-coating role on the oxidation kinetics of a pyritic sludge doped with fly ash. Geochimica et Cosmochimica Acta 71(8), 1921-1934. https://doi.org/10.1016/j.gca.2007.01.019.
Pérez-López, R., Cama, J., Nieto, J. M., Ayora, C., & Saaltink, M. W. (2009). Attenuation of pyrite oxidation with a fly ash pre-barrier: Reactive transport modelling of column experiments. Applied Geochemistry 24(9), 1712-1723. https://doi.org/10.1016/j.apgeochem.2009.05.001.
Roy, V., Demers, I., & Plante, B. (2020). Kinetic Testing for Oxidation Acceleration and Passivation of Sulfides in Waste Rock Piles to Reduce Contaminated Neutral Drainage Generation Potential. Mine Water Environment 39, 242-255. https://doi.org/10.1007/s10230-020-00680-z.
Sahoo, P., Kim, K., Equeenuddin, S., & Powell, M. (2013). Current approaches for mitigating acid mine drainage. Rev Environ Contam Toxicol, 226, 1-32. https://doi.org/10.1007/978-1-4614-6898-1_1.
Shi, M., Min, X., Ke, Y., Lin, Z., Yang, Z., Wang, S., . . . Wei, Y. (2021). Recent progress in understanding the mechanism of heavy metals retention by iron (oxyhydr)oxides. Science of The Total Environment 752, 141930. https://doi.org/10.1016/j.scitotenv.2020.141930.
Skousen, J. R., Geidel, G., Foreman, J., Evans, R., & Hellier, W. (1998). Handbook of technologies for avoidance and remediation, p 131. Morgantown: National Mine Land Reclamation Center.
Skousen, J., Zipper, C., & Rose, A. (2017). Review of Passive Systems for Acid Mine Drainage Treatment. Mine Water Environ 36, 133–153. https://doi.org/10.1007/s10230-016-0417-1.
Swedlund, P., Miskelly, G., & McQuillan, A. (2009). An attenuated total reflectance IR study of silicic acid adsorbed onto a ferric oxyhydroxide surface. Geochimica et Cosmochimica Acta 73, 4199-4214. https://doi.org/10.1016/j.gca.2009.04.007.
Tu, Z., Wu, Q., He, H., Zhou, S., Liu, J., He, H., . . . Reinfelder, J. R. (2022). Reduction of acid mine drainage by passivation of pyrite surfaces: A review. Science of The Total Environment 832, 155116. https://doi.org/10.1016/j.scitotenv.2022.155116.
Van den Eynde, V. C., Paradelo, R., & Monterroso, C. (2009). Passivation techniques to prevent corrosion of iron sulphides in roofing slates. Corrosion Science 51(10), 2387-2392. https://doi.org/10.1016/j.corsci.2009.06.025.
Vandiviere, M., & Evangelou, V. (1998). Comparative testing between conventional and microencapsulation approaches in controlling pyrite oxidation. Journal of Geochemical Exploration 64(1), 161-176. https://doi.org/10.1016/S0375-6742(98)00030-2.
Vempati, R., Loeppert, R., Dufner, D., & Cocke, D. (1990). X-ray photoelectron spectroscopy as a tool to differentiate silicon-bonding state in amorphous iron-oxides. Soil Science Society of America 54, 695-698. https://doi.org/10.2136/sssaj1990.03615995005400030010x.
Wang, S., Zhao, Y., & Li, S. (2019). Silicic protective surface films for pyrite oxidation suppression to control acid mine drainage at the source. . Environ Sci Pollut Res 26, 25725–25732. https://doi.org/10.1007/s11356-019-05803-w.
Zhang, T., Zhang, C., & Du, S. (2023). A review: The formation, prevention, and remediation of acid mine drainage. Environmental Science and Pollution Research 30, 111871–111890. https://doi.org/10.1007/s11356-023-30220-5.
Zhang, Y., & Evangelou, V. (1998). Formation of ferric hydroxide-silica coatings on pyrite and its oxidation behaviour. Soil Science, 163, 53-62. http://dx.doi.org/10.1097/00010694-199801000-00008.

Download PDF

Reviewers agreed at journal
24 Oct, 2024
Reviewers invited by journal
24 Oct, 2024
Editor assigned by journal
09 Oct, 2024
First submitted to journal
08 Oct, 2024

You are reading this latest preprint version

Prevention of Acid Rock Drainage formation through pyrite inhibition by silica coating

Status:

Version 1

Abstract

Figures

Introduction

Materials and methods

Waste rock characteristics and analyses

Leaching experiment

Experimental set up

Leachate sampling and analyses

Results and discussion

Waste rock characteristics

Silica precipitates

Leaching characteristics

Prevention of ARD formation by silica treatment in waste rock

Conclusions

Declarations

Ethical approval

Consent to participate

Consent for publication

Competing interests

Funding

Author Contribution

Acknowledgements

Data availability

References

Supplementary Files

Status:

Version 1