Statement of Novelty

This study investigates the feasibility of metal recovery from mine tailings using a hydrometallurgical process with the addition of biochar–Fe composites. Optimization of experimental conditions can improve the recovery of zinc and copper while minimizing arsenic mobility. The use of biochar–Fe composites reduced arsenic leaching, promoting its immobilization in the final residue. In addition, leaching of mine waste in the presence of biochar–Fe composite led to residues with phytostimulation properties, contributing to reduce their environmental impact.

Introduction

Mining and processing of mineral resources for metals production generate large amounts of waste. Initially, these wastes were discharged into the environment, in the absence of any previous treatment, but, in recent years, these wastes have been deposited in controlled tailings dams. However, failures in tailings dam storage have originated not only several environmental problems [1], but also multiple human risks. One of the most famous accidents in Europe was the spill of large-scale sulfide tailings dam in Los Frailes (Aznalcóllar, Spain). It caused the inundation of the floodplains of the rivers Agrio and Guadiamar [2, 3]. Furthermore, the tailings flood was on the verge to affect the Guadalquivir marshlands, threatening the Doñana Natural and National Parks’ ecosystems. In order to prevent similar issues in the future, there is a great need to perform new economically feasible ways of mining waste management that not only favor the valorisation of available resources, but also, to perform a safe environmental remediation recovering lands for other uses. According to article 5 of the Directive 2006/21/EC related to Extractive Waste (EWD), the management plan shall have the objective of (a) the prevention or reduction of extractive waste generation and of its harmfulness; (b) the recovery of extractive waste by means of recycling, re-using, or reclaiming such waste; and (c) ensuring the short- and long-term safety of the extractive waste, in particular by considering as part of the design [4].

Some extractive mine wastes still contain valuable and/or critical metals and their recovery is one of the most effective techniques to prevent metals pollution, achieving an economic advantage [5,6,7,8]. However, in most cases, the low concentration of valuable metals hinders their economic benefit. It is well known that hydrometallurgy is one of the most efficient technologies to recover valuable metals from low-grade ores and wastes, being sulfuric acid the most often used leaching system due to low price, availability, and dissolution capacity [9,10,11]. However, some mining wastes related to the exploitation of metallic deposits like copper, zinc, or lead are rich in sulfide minerals and show low solubility in sulfuric medium [12]. Recently, different researchers indicated that some carbon materials such as black carbon or activated carbon improved the kinetics of leaching of sulfide minerals such as chalcopyrite [13, 14] or enargite [15, 16]. The presence of carbon materials in the adequate ratio [14] leads to a substantial increase in copper extraction rate, probably due to the decreases in the redox potential, as well as galvanic interaction between sulfide particles and the carbon material surface [17, 18].

Old mine wastes, depending on the geological context of metal deposits, contain significant residual sulfide and sulfosalt content, with different reactivity under atmospheric conditions that are responsible for generating contamination. Previous works performed by our research group have shown that activated carbon and charcoal addition can also increase the extraction of Zn and Cu from mine tailings. In spite of the different properties between activated carbon and charcoal, their effect in the leaching of Zn and Cu was similar [19]. This result can open the prospect to use low-cost carbon materials, like biochar, as catalysts for the leaching of metals and the development of new hydrometallurgical processes. In addition, utilization of biomass, particularly biomass waste, to produce biochar can result in a change to a more circular economy. The potential use of carbon materials as catalysts in the leaching of low-grade ores or old mine tailing with altered minerals could have several advantages. For example, the great stability of carbon materials facilitates its recovery in the process and re-utilization, in a similar way to the use of activated carbon in gold metallurgy [20]. On the other hand, the presence of carbon materials, like biochars, in the final waste could limit their toxicity and environmental impact [21]. Currently, studies focused on the use of carbon materials during leaching processes for metals recovery from mining wastes are insufficient. For this reason, the aim of this work is to study the addition of two biochar-Fe composites in the leaching of zinc and copper from a mining waste. The arsenic concentration in the leaching systems was also analyzed. Finally, a phytotoxicity test was performed on the final residue obtained after leaching experiments.

Materials and Methods

Materials Selection and Characterization

One mining waste sample (MW) was selected from one abandoned mining deposit located in the southeast of Spain. The origin of MW was an abandoned zinc/lead ore extraction mine. Sample was air-dried, crushed, and sieved under 50 µm, using a ceramic mill. Wavelength X-ray fluorescence (WDXRF) was achieved in an ARL ADVANT′XP+ sequential model from THERMO (SCAI-Malaga University) and metal composition was obtained by an UNIQUANT Integrated Software. A Bruker diffractometer, model D8 Advance A25 (SCAI-Malaga University), was used for X-ray Diffraction (XRD) analysis.

Two biochar–Fe composites obtained by pyrolysis of pruning waste impregnated with 5 wt% ferric sulfate (BM–Fe) or pyrolysis of hydrochar from pruning waste impregnated with 5 wt% ferric sulfate (HM–Fe) were used. Ingelia (Náquera, Spain) supplied pruning waste and hydrochar from pruning waste. The two types of biochar–Fe composites were prepared by impregnation with ferric sulfate salt of biomass waste or corresponding hydrochar, followed by pyrolysis at 500 °C for 5 h as was described in detail by Álvarez et al. [22].

The two biochar–Fe composites were air-dried, crushed, and sieved below 100 µm using a ceramic mill. Characterization of samples was performed according to the following properties: pH and redox potential (Eh) were determined with a biochar/distilled water ratio of 0.1/25 (g mL−1), using a Crison micro pH 2000 and Eh in a pH 60 DHS, respectively. Elemental analysis (C, H, N, O and S in %) was performed using a LECO CHNS 932 Analyser by dry combustion. Ash content (%) was calculated by combustion of samples at 850 °C in a Labsys Setaram TGA analyzer. Twenty mg of each sample was heated at a rate of 15 ℃/min up to 850 ℃ using 30 mL min−1 of air. Oxygen was obtained by difference as 100%-(%C + %H + %N + %S + %Ash). Following that, O/C and H/C ratios were calculated from the elemental analysis results. Porosity (%) was determined by Hg porosimetry, which was carried out using a Micromeritics AutoPore IV 9500 equipment. BET surface area (SBET), was analyzed by nitrogen adsorption isotherm, which was carried out at 77 K in a Micromeritics Tristar 3.00.

Leaching Experiments

Leaching experiments were performed using a thermostatic bath with stirring model GFL 1083 (heating power of 1500 W and the voltage 230 V). The temperature conditions and stirring speed were 90 °C and 250 rpm, respectively. Approximately, 2.5 g of MW were weighed in a 250 mL borosilicate glass jar. Then, 50 mL of leaching agents (H2SO4 0.25 M or 0.17 M solutions with pH value of 0.6 and 1.0 respectively). Sulfuric acid with 95–98% purity was supplied by Sigma-Aldrich. Except for the control samples, each carbon material was added to MW in a 1/0.5 or 1/0.25 ratio (w/w).

One mL of each sample of the supernatant liquor was withdrawn at different reaction times (1, 2, 4 and 6 h). The sampling procedure was as follows: Firstly, in order to let the sample settle and favor the decantation of solids, the stirring was stopped. After that, 1 mL of the supernatant solution was removed, filtered, and transferred to a 25 mL graduated flask, making up to volume with distilled water. In order to compensate the extracted leaching solution and maintain the same conditions throughout the system, 1 mL of the corresponding sulfuric acid solution was added. After 6 h of leaching, the stirring was stopped, and the supernatant was allowed to cool down. After that, the pulp was filtered and the solid was washed twice with 50 mL of H2SO4 solution of pH 2. pH and Eh of leaching solution were determined along leaching experiments using a Crison micro pH 2000 and a pH 60 DHS, respectively.

Zn, Cu, As, and Sb concentration in the leaching and washed solutions were determined using an inductively coupled plasma mass spectrometry, model ICP-MS Elan DRCe (SCIEX Perkin Elmer) from SCAI-Malaga University.

Phytotoxicity of Residue After Leaching Experiments

The potential phytotoxicity of the final residue generated after leaching experiments was determined using the germination test described by Zucconi et al. [23]. Briefly, five seeds of Lepidium sativum were placed on petri dishes with filter papers at the bottom, then 5 mL of aqueous extract (1/10 w/v) of each sample was added. Seeds were maintained in the dark at 28 °C. Germination percentages (G) with respect to control (distilled water) and root lengths were determined after 48 h. The germination index (GI) was estimated as GI = G·Le/Lc where G is the percentage of germinated seeds in each extract with respect to the control, Le is the mean total root length of the germinated seeds in each extract, and Lc is the average root length of the control.

Statistical Analysis

The significance of the differences among means was assessed by analysis of variance (ANOVA), using the Tukey test as a post hoc. Every analysis was performed in triplicate (n = 3).

Results and Discussion

Characteristics of Samples

Analysis by Rietveld method of XRD (Fig. 1) shows that main crystalline mineral species present in MW sample were the following: muscovite (49.6%), followed by quartz (38.9%), corkite (6.9%), and calcite magnesium (3.5%) The content of litharge stannite and calcite was lower than 1%. Table 1 summarizes chemical composition of MW, the background ranges, and generic reference levels of some trace elements in the natural soils of the Region of Murcia (Spain) [24]. Background levels are defined as the natural content of metals in a soil from a determined area without any anthropogenic influence. On the other hand, generic reference levels indicate the concentration of metals in the soil which do not pose a greater risk than the maximum acceptable to human health or ecosystems. According to that, each country or region determines the regulatory standards over which a soil can be considered contaminated for different land uses, such as agricultural, industrial, and urban. Although contents of Cu (0.0435% ± 0.0022), As (0.151% ± 0.0075), Co (0.0079% ± 0.0007), Zn (1.38% ± 0.05), and Pb (2.18% ± 0.07) of MW were lower than exploitable concentrations in ores, MW would can be classified as a contaminated soil as the concentration of Zn, As, Cu, Co, and Sb exceeded the background and generic reference levels. These metals are listed as very toxic trace elements due to their high degree of mobilization in soils, and the US Environmental Protection Agency [25] registers most of them as priority pollutants.

Fig. 1
figure 1

XRD pattern of MW sample

Full size image
Table 1 Chemical composition of MW sample compared to ranges of some trace elements in soils of the Region of Murcia compared to the abnormal values from the mine wastes of the Sierra Minera [24]
Full size table

Table 2 shows the main properties of two biochar–Fe composites used in this work. HM–Fe showed high C content than BM–Fe. The lowest H/C ratios and, consequently, the highest aromaticity corresponded to BM–Fe (0.17), whereas HM–Fe showed the highest content of oxygen functional groups. The highest surface area corresponded to HM–Fe and the lowest to BW–Fe. Finally, two samples show basic pH and low Eh (mV) values.

Table 2 Main characteristics of biochar–Fe composites
Full size table

Percentage of Metals Recovery

Figure 2a provides the recovery of Zn (%) from MW treated with sulfuric solutions at pH 0.6 and 1.0, after 6 h of leaching experiments. Leaching experiments were performed at 90 °C in order to increase the kinetic of the reactions [9]. The percentage of total Zn recovered is higher in sulfuric solutions with pH 0.6 (varying from 72.8 to 76.3%) than in solution with pH of 1.0 (ranging from 63.5 to 70.7%). In general, the addition of biochar-Fe composites increased the total Zn recovered at pH 1.0, whereas at pH 0.6, the addition of biochar-Fe composites slightly diminished the amount of Zn recovered.

Fig. 2
figure 2

Recovery of Zn (a) and Cu (b)

Full size image

The addition of BM–Fe with ratio 1/0.25 and using the leaching solution at pH 0.6, significantly increased the recovery of Cu (Fig. 2b). In a similar way to the recovery of Zn, using sulfuric leaching solution at pH 1, the addition of both biochar-Fe composites increased the recovered of Cu with respect to control.

With respect to As (Fig. 3a), the highest leaching was obtained at low pH (0.6). It is important to highlight that the addition of BM–Fe and HM–Fe biochars significantly reduced the leaching of As. During carbon-catalyzed atmospheric leaching of enargite, Jahromi et al. [26] concluded that the presence of activated carbon with high ferric concentration can immobilize As, which precipitates as scorodite particles. In our research, leaching was performed without ferric acid addition. However, biochars were prepared with Fe in their composition leading to higher As immobilization in the final residue. Previous works have proven that Fe-impregnated biochar has considerable ability for As immobilization in contaminated soils, decreasing the available As and increasing the As bound to Fe oxides of biochars [27]. The decrease in the leaching of As was an important improvement in the recovery of metals from mining wastes or minerals as As could be immobilized in the residue, decreasing costs of leaching solution purification. The behavior of Sb was similar to that of As, decreasing the amount of Sb leached with the addition of BM–Fe and HM–Fe (Fig. 3b). Carbon structure and the presence of Fe oxides in their surface can improve their Sb adsorption capacity [28,29,30]. As Tighe et al. [31] indicated, similarly to As, Sb can be immobilized in the residue due to their strong affinity to non-crystalline Al and Fe hydroxides.

Fig. 3
figure 3

Recovery of As (a) and Sb (b)

Full size image

Evolution of Zn, Cu, and As Extraction During Leaching Experiments

The evolution of Zn, Cu, and As concentration in the leaching solution is represented in Figs. 4, 5, and 6, respectively. A similar trend over time was observed for three elements. Cu, Zn, and As rise steadily between 1 and 4 h of leaching, increasing significantly between 4 and 6 h.

Fig. 4
figure 4

Extraction of Zn (%) in the leaching solution at pH 0.6 (a, b) and pH 1.0 (c, d)

Full size image
Fig. 5
figure 5

Extraction of Cu (%) in the leaching solution at pH 0.6 (a, b) and pH 1.0 (c, d)

Full size image
Fig. 6
figure 6

Extraction of As (%) in the leaching solution at pH 0.6 (a, b) and pH 1.0 (c, d)

Full size image

Eh and pH Evolution During Leaching Experiments

Figure 7 shows the Eh evolution during leaching of MW and MW treated with BM–Fe and HM–Fe H2SO4 solution with pH 0.6 (Fig. 6a) and pH 1.0 (Fig. 6b). In general, the addition of biochars decreased the Eh of leaching systems. This result was similar to previous results obtained by Álvarez et al. [19]. Other researches have observed Eh reduction after carbon black addition to chalcopyrite in the sulfuric acid leaching system [17]. Galvanic interactions may occur between carbon structures and minerals in the acidic medium, which may decrease the Eh of the system [32]. It is expected that the addition of carbon materials lead to a decrease in Eh due to their C content and, consequently, low Eh values (Table 1). At pH 0.6, the main reductions in the Eh corresponded to samples MW/BM–Fe and MW/HM–Fe with ratio 1/0.5. However, at pH 1.0, the Eh was similar using ratios 1/0.5 and 1/0.25.

Fig. 7
figure 7

Eh (mV) evolution of samples in leaching solutions with pH 0.6 (a, b) and pH 1.0 (c, d)

Full size image

Figure 8 provides the pH evolution during leaching of MW and MW treated with BM–Fe and HM–Fe in H2SO4 solution with an initial pH of 0.6 (Fig. 7a and b) and with a pH of 1.0 (Fig. 7c and d). In general, the pH of the leaching systems increased with time (from 0 to 6 h). However, it is important to note that when leaching is performed in the presence of biochar–Fe composites and, especially with more concentrated H2SO4 solution, the pH increment was lower than without biochar addition (Fig. 8a and b). This result can be due to the modification of biochar surface with leaching agent and the generation of acidic groups on their surface, with a low acid consumption during leaching process. Álvarez et al. [18] observed that the addition of carbon materials significantly decreased the pH of the ferric acid solution (Fe3+/H2SO4 0.5 M).

Fig. 8
figure 8

pH evolution of samples in leaching solutions with pH 0.6 (a, b) and pH 1.0 (c, d)

Full size image

Phytotoxicity of Residue Produced After Leaching Process

Table 3 shows the germination index of residues obtained after leaching and washing of MW and MW treated with BM–Fe and HM–Fe. In order to compare, the germination index of original MW sample before leaching was also determined (MW control). According to Emino and Warman [33], samples with GI > 100% can be considered phytonutrient or phytostimulant. In this study, the addition of BM–Fe increases the GI of final leaching residue from 80% in the MW control (and 71–77% in MW after leaching) to values higher than 100%. With the addition of HM–Fe, the GI of final residues was different depending on the pH used during the leaching process, being lower after the use of leaching solutions with pH 0.6.

Table 3 Germination index (GI), pH, and electrical conductivity (EC) of MW before and after leaching processes
Full size table

With respect to electrical conductivity (EC), of the EC of MW significantly decreased after leaching and washing experiments, from 2.95 dSm−1 to values ranging from 0.05 to 0.40 dSm−1.

Results summarized in Table 3 showed that leaching of mine waste with the addition of BM–Fe in 1/0.5 or 1/0.25 ratios led to residues without phytotoxic characteristics. In fact, in these cases, the final residues showed phytoestimulation properties (GI > 100) indicating the potential used as growing media for the development of phytoremediation technologies in these mining areas.

Conclusions

The addition of biochar–Fe composites as catalysts in the leaching of metals from mine wastes could be a promising alternative to traditional hydrometallurgical processes. The use of biochars BM–Fe and HM–Fe did not improve the amount of zinc recovered, but was able to reduce, significantly, the arsenic leaching, promoting its immobilization in the final residue. The addition of biochar BM–Fe in a low ratio (1/0.25), using leaching solution of pH 0.6, increased copper recovery. Leaching of mine waste in the presence of BM–Fe generates residues with phytostimulation characteristics. The use of biochar BM–Fe as catalyst in hydrometallurgical processes can open an interesting line of research, as the final residues can be used as growing media for phytoremediation technologies, reducing their environmental impact. Further research will be necessary to optimize the characteristics of the biochars and their use in the leaching of metals.