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Article

Stabilization of As and Heavy Metal-Contaminated Soils by Two Mine Drainage-Treated Sludges

by
Hyunji Tak
1,
Seonhee Kim
2,
Kyeongtae Kim
3,
Sookyun Wang
4 and
Minhee Lee
3,*
1
Marine Environment Research Division, National Institute of Fisheries Science, Busan 46083, Republic of Korea
2
Geoanalysis Center, Geology Division, Korea Institute of Geoscience and Mineral Resources, Daejeon 34132, Republic of Korea
3
Major of Environmental Geosciences, Pukyong National University, Busan 48513, Republic of Korea
4
Department of Energy and Resources Engineering, Pukyong National University, Busan 48513, Republic of Korea
*
Author to whom correspondence should be addressed.
Minerals 2023, 13(2), 148; https://doi.org/10.3390/min13020148
Submission received: 21 November 2022 / Revised: 7 January 2023 / Accepted: 16 January 2023 / Published: 19 January 2023
(This article belongs to the Special Issue Groundwater Protection Challenges in Mine: Retention of Hazardous Substances with Solidification/Stabilization Implementation)
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Abstract

:
This study explored and analyzed the potential of the practical use of acid mine drainage-treated sludge (AMDS) as a new soil stabilizer for arsenic (As) and heavy metals. Various analyses, toxicity evaluations, and extraction batch experiments were performed to investigate the characteristics of the AMDS as a soil stabilizer and to identify the main mechanisms to fix As and heavy metals on the AMDS in soil. Two types of AMDS, copper metal mine drainage-treated sludge (MMDS) and coal mine drainage-treated sludge (CMDS) and four contaminated soils with different pollution scenarios were used in the experiments. ‘Soil A’ and ‘Soil D’ were mainly contaminated with Cd, Pb and Zn. ‘Soil B’ and ‘Soil C’ were contaminated with As. Results from XRD, XRF, SEM-EDS, TG-DTA, and BET analyses suggested that AMDS is mainly composed of Fe- and Ca- bearing minerals such as CaCO3, Ca(OH)2 and amorphous Fe-oxide (hydroxide), which have a large surface area and high adsorption capacity for As and heavy metals. From batch extraction experiments, the Pb stabilization efficiency of both of the AMDSs in soil A, which has a high Pb and Zn content, was higher than 90%. The high heavy metal stabilization efficiency comes directly from the electrostatic attraction between metal cations and the negatively charged AMDS surface and/or from the co-precipitation of metal oxide (hydroxide) and CaCO3, which occurs comprehensively on the AMDS surface. In the case of Zn, the stabilization efficiency in soil A was somewhat low due to the adsorption competition with Pb, but the Zn stabilization efficiency of the CMDS in soil A was higher than 80% (70% or higher for the MMDS). For soil D, the Zn stabilization efficiency of two AMDSs was higher than 85% because of the lower concentration of other heavy metals in soil D, compared to in soil A. The As stabilization efficiency of the AMDSs in soil contaminated with As (soil B and soil C) was higher than 85%, (mostly > 95%). The overall stabilization efficiency of two AMDSs for heavy metals and As were higher than 75% and 85% (mostly > 90%), respectively, regardless of soil type. We concluded that this high As stabilization efficiency was due to the formation of a new complex by ligand exchange between the Fe- (oxide) hydroxide and the arsenate and also to the cation bridge effect between the AMDS surface and the arsenate as well as the co-precipitation.

1. Introduction

There have existed about 2000 metal mines and 400 of coal mines in South Korea, and more than 90% of them have been closed, and mine tailings and rock waste have piled up in ruin or scattered without control for a long time [1,2]. The oxidation/reduction process of mine tailings and rock waste in abandoned mines have continuously generated much acid mine drainage (AMD), leaching out many heavy metals at a low pH level and deteriorating the quality of groundwater and surface water [3]. The proper treatment of the AMD before it is drained into the nearby water systems is critical, and more than 50 treatment facilities for AMD with high heavy metal content have operated in South Korea [4]. Chemical treatment processes, including the pH-neutralization and the precipitation, have been widely used to treat AMD [5,6,7]. Among them the pH titration–precipitation process, which uses Ca(OH)2 and CaCO3, is one of the most used methods for its the cost and the operation simplicity [8,9]. In this treatment, the granulated Ca(OH)2 and CaCO3 are mixed with AMD to neutralize the low pH of the AMD and induce co-precipitation to separate heavy metals, including Fe, from the AMD [10]. There also exists the treatment of acid mine tailings by using the alkaline and cementitious amendments, fundamentally preventing the AMD generation [11,12]. More than 5000 tons of sludge as the byproduct after AMD treatment for coal mines and metal mines have been annually generated in South Korea. It was classified as industrial waste regardless of the composition and toxicity, and most of it was safely disposed of in landfill sites. Every province of South Korea has suffered from the limitations of the landfill sites, and these days, how to recycle the acid mine drainage-treated sludge (ADMS) has emerged as an important research topic [13,14,15,16].
Many studies have been conducted to remove heavy metals and arsenic (As) from soils: soil washing, electrokinetic extraction, phytoremediation, and stabilization process [17,18,19]. The stabilization process is one of main processes for controlling contaminated soils by decreasing the extraction or leaching rate of heavy metals and As through immobilizing amendments [20,21,22]. The use of proper amendments to accelerate immobilization processes, such as sorption, precipitation and complexation reactions can induce soils to reduce mobility and bioavailability of heavy metals and As [23,24,25,26,27]. One of the cheapest and the most effective materials for soil stabilization is lime (CaO) and lime-based stabilization, which is an effective remediation alternative to immobilize As or other heavy metals in contaminated soils [28,29,30,31]. However, the pH increase according to the addition of lime to soil is seriously limited for using lime as a stabilization amendment in the field. As an alternative to conventional lime, stabilizers such as limestone, glass fiber, coal fly ash, gypsum, oyster shell, steel slag, and Mn-Fe oxide have been studied for heavy metals and/or As-contaminated soil [22,32,33,34,35,36,37,38,39,40].
According to the remediation cases for heavy metal and As-contaminated sites in South Korea, the soil washing method has been used at the highest rate of 61.6% (165/268), followed by electrokinetic separation (48/268), chemical treatment (32/268), and phytoremediation (14/268) over the last three decades [41]. In contrast, the solidification/stabilization method has been used for 25% of the selected remedies in the United States and other methods were used less than 10% [42,43]. Until now, the domestic stabilization process in South Korea was implemented locally on heavy metal-polluted farmland near abandoned mines in accordance with the law on the protection against harm in mines (Article 11, the Law about Mine Damage Prevention and Recovery) [44], but the application of the stabilization method in South Korea has been very limited. However, the enforcement ordinance of the Soil Environment Conservation Act was amended so the stabilization method could be applied to extensive sites where it is difficult to remove the pollutants thoroughly. The revision of the law is expected to expedite the development of appropriate stabilization technologies based on the site characteristics and pollution situations.
Because acid mine drainage-treated sludge (AMDS) is a by-product generated from the treatment process of AMD for pH-neutralization and heavy metal removal, its usage as a soil stabilizer needs to take into account not only the composition but also the leaching rate of toxic components such as heavy metals and As over a long period. However, only a few studies have been performed on the properties of AMDS for recycling, and its use for stabilizing heavy metals and As in soils has been studied at the very first step [45,46]. A clear understanding of the various reactions among AMDS, soil particles and water is essential for selecting the AMDS most suitable for use with a specific soil. However, studies on the reaction mechanisms involving AMDS in soil are also very limited.
This study focused on characterizing the AMDS to reduce the extraction of heavy metals and As from contaminated soil and on evaluating the feasibility of the AMDS as a soil stabilizer by investigating the main stabilizing mechanism of AMDS. In this paper, we also quantitatively present how to increase the stabilization efficiency in various soils contaminated with heavy metals and As using the different characteristics of two AMDSs, AMDS originating from coal mines (CMDS) and AMDS originating from copper metal mines (MMDS). Batch experiments were performed to investigate the efficiency of the AMDS as immobilizing amendments to reduce the heavy metal and As leaching from the contaminated soil. Pilot scale column as a physical model for a genuine contaminated soil environment was designed and the As leaching rate change using AMDS in diverse treatment conditions of the column experiment was measured. These results provide insights not only into the use of AMDS as a stabilizer for heavy metal and As-contaminated soils, but also for finding other recycling purposes for AMDS.

2. Material Preparation

2.1. Preparation of Soils

For this study, four domestic heavy metal and As-contaminated sites were selected based on different pollution scenarios (‘Soil A’: metal waste landfill soil in a military unit, ‘Soil B’: top soil at a mine tailing storage site, ‘Soil C’: farmland soil around an abandoned mine, and ‘Soil D’: sloped land soil located near an explosives disposal military site). ‘Soil A’ and ‘Soil D’ were mainly contaminated with cadmium (Cd), lead (Pb) and zinc (Zn). ‘Soil B’ and ‘Soil C’ were contaminated with arsenic (As). Figure 1 shows the locations of the four soil and two AMDS sampling sites in South Korea. Fifty kilograms of the soil sample at each site were collected at a 10–30 cm depth. The soil samples were dried in an oven at 30 °C for 2–4 days and sieved at 2 mm in diameter for the experiments. The aqua regia extraction method (ISO 11466 method) was applied to extract the heavy metal and As contents from soil samples, and an inductively coupled plasma-optical emission spectrometer (ICP-OES; Optima 7000DV, Perkin Elmer, MA, USA) and an inductively coupled plasma-mass spectrometer (ICP-MS; NexION 300D, Perkin Elmer, MA, USA) were used to measure their concentrations [47]. All sampling and analytical processes for soils were followed. The pH of the soils was measured using a pH meter (Orion Star-A211, Thermo Fisher Scientific, MA, USA) based on the standard method in South Korea. All sample preparation and analyses for soil and AMDS proceeded according to these standard manuals in the notification by regulation [48].

2.2. Preparation of AMDSs

To investigate the stabilization efficiency of the AMDS for heavy metals and As, two types of AMDS in South Korea were used as the stabilizer in this study according to their different AMD origins and mining material (coal and metals). One sample of the AMDS was obtained after the treatment of the AMD, which originated from a metal abandoned mine (MMDS). The MMDS used in this study was collected from the Ilgwang mine water treatment facility located in Ilkwang-myeon, Gijang-gun, Busan. The main mining metal of the Ilkwang mine was copper, and in addition, gold and silver were partially produced; the mine closed down in the 1990s [49]. About 5 tons of AMD drained out from the mine head per day and it was reported that the concentration of heavy metals of the inflow AMD into the water treatment facility exceeded the Korean drainage tolerance limit (Table 1) [50]. In the Ilkwang mine water treatment facility, the AMD was treated using a semi-active treatment method, including a pH neutralization process and a co-precipitation process, using agents such as lime (CaO), limestone (CaCO3), slaked lime (Ca(OH)2), and caustic soda (NaOH), which generated many sludge byproducts (MMDS) on the bottom of the precipitation tank. The AMD treatment process at the Ilkwang mine is shown in Figure 2a.
The other type of sludge was AMDS that was produced at an abandoned coal mine (CMDS) in Jeongseoun-gun, Kangwondo. The abandoned Hambaek coal mine had been active for about 20 years and about 30,000 tons of mine tailings and waste rock fragments had been left at the storage site, generating 5–10 tons of AMD daily. The AMD originating from the abandoned Hambaek mine was treated using the electro-purification treatment method, including electrolysis reaction and co-precipitation processes. The electro-purification treatment method combines heavy metals dissolved in the AMD with OH- generated during the electrolysis, precipitating them as solids [51] (Figure 2b). When electric charges are supplied from the electrolysis tank, a large amount of metal oxide (or hydroxide) is generated through an oxidation–reduction reaction, which accelerates the co-precipitation of metals in the form of CMDS on the bottom of the settling tank.

3. Experimental Methods

3.1. Chemical Characterization of AMDSs

Two types of AMDS were collected from the bottom of each artificial fen at the AMD treatment facility of the abandoned Hambaek coal mine and the abandoned Ilkwang metal mine. After the organic debris and rock fragments were removed from the AMDS at each fen, they were naturally dried indoors at room temperature. They were pulverized for the analysis for components using XRF and the pH of each was measured using a pH meter. To verify the leaching safety of toxic compounds from the AMDS, two toxicity leaching tests were also performed to evaluate the risk that toxic heavy metals and As might be leached from them. The toxicity characteristic leaching procedure (TCLP) test for heavy metals and As was conducted according to the US-EPA 1311 method [52]. The synthetic precipitation leaching procedure (SPLP) test was also conducted according to the US-EPA 1312 method [53]. The SPLP test was used to estimate the amount of outflow of heavy metals and As in acid-rain conditions, and the mixture of dilute nitric acid and sulfuric acid was used as an extraction solution (pH 4.2 ± 0.05). To identify the total amount of heavy metals and As in the AMDS, they were also extracted using the aqua regia extraction method and their concentrations in each AMDS were analyzed on the ICP-OES (ICP-MS for As) [48].

3.2. Mineralogical and Structural Characterization of AMDSs

To identify the main mechanisms of the AMDS that cause As or heavy metal stabilization in soils, various analyses were performed. The mineralogical properties of the AMDS were analyzed using an X-ray diffractometer (XRD; X’Pert3 Powder, Malvern Panalytical, Malvern, United Kingdom) using Cu Kα radiation (λ = 1.54060 Å) over a 2θ range of 10–80° and X-ray power of 40 kV/30 mA at a scan. The principal components of the AMDS were also analyzed by an X-ray fluorescence spectrometry (XRF; XRF-1800, Shimazu, Kyoto, Japan). The powdered samples of each AMDS were prepared as pellets on a boric acid support. The measurements were performed using a Shimadzu (XRF-1800) sequential spectrometer equipped with a Rh Kα X-ray tube operated at 40 kV and 95 mA. The XRF spectrum analysis used a type of calibration method known as fundamental parameters (FP). The surface area of each AMDS was calculated by fitting the released amount of N2 gas on the BET isotherm curve (a function of the relative pressure increase) [54]. The BET analysis model used in the analysis was ASAP 2420 manufactured by Micromeritics Instrument in USA. It was generally agreed that the more Ca-hydroxide or CaCO3 components in the AMDS, the higher the possibility of stabilizing the As and heavy metals [55]. The existence of specific molecules in the material can be determined through the mass weight changes caused by temperature increases. Therefore, the thermo gravimetric-differential thermal analyzer (TG-DTA; DTG-60H, Shimazu, Kyoto, Japan) analysis was performed on the AMDS. From only the result of the TG analysis, it was difficult to clearly find the detailed temperature range where the weight change appeared, so the DTG (derivative thermo gravimetry) curve using the differentiated value of the TG curve was also used in this study and the weight reduction rate was calculated by measuring the area of the DTG peak. Through the TG-DTA analysis, the weight reduction in specific molecules according to the temperature increase was identified, and the existence of hydrous and carbonate materials in the MMDS and the CMDS was investigated, and the TG-DTA results were compared with their XRD results. The scanning electron microscope-energy dispersive spectrometer (SEM-EDS; VEGA II LSU, TESCAN, Brno, Czech Republic) analysis for the AMDS was also performed to visualize the surface structures of the MMDS and the CMDS, and by the EDS analysis, the major components of each AMDS were additionally identified. To observe the surface structure of the AMDSs, prepared samples were spread onto carbon tape and mounted on an aluminum specimen holder. Then, powdered AMDS specimens sputtered coated using platinum (Pt) for SEM-EDS analyses. The results were compared to those of the XRF results. To ensure the reliability of the experimental data, the experimental measurements were repeated three times and their arithmetic mean was determined as the final value for each measurement.

3.3. Batch Extraction Experiments for the Stabilization Efficiency of AMDSs

The stabilization efficiency of the AMDS was evaluated from how much the amount of heavy metals and As extracted from soils decreased after the AMDS was added to soil. Batch extraction experiments were performed to investigate the reduction in the amount of heavy metals and As extracted from the soil caused by two types of AMDS. Soil samples were taken from four sites (Soil A, Soil B, Soil C and Soil D) for the batch experiments. For the batch extraction experiments, various extraction conditions were applied to determine the parameters affecting the extraction reduction in heavy metals in the AMDS. To determine the extraction reduction capacity of the AMDS for As and heavy metals, different amounts of each type of AMDS were added to the soil in the experiment. Dried contaminated soil, sieved with No. 10 mesh (2 mm in diameter) was mixed with powdered AMDS (CMDS or MMDS) at various ratios (0, 1, 3, 5, and 7 wt.% of soil). A total of 40 g of soil + AMDS was immersed in 120 mL of distilled water (titrated at pH 6) in a 200 mL Teflon capped glass tube. The tube was shaken at 20 °C and 150 rpm in the thermo-hygrostat for 2 h and then left stationary for 8 h. The fixation of As or heavy metal on the AMDS occurred for an additional stabilization time (0, 12 h, 36 h and 60 h), and a supernatant of mixed material was separated from the mixture solution and then filtered with a 5B filter. The heavy metal and As concentrations in the filtered solution were analyzed using ICP-OES (ICP-MS for As). After taking the supernatant 60 h after fixation, the pH of the remaining supernatant was also measured to confirm the pH change of the extraction solution. More details about the batch experiment can be found in the previous study [56]. The efficiency of extraction reduction (the stabilization efficiency) by the AMDS was calculated using Equation (1).
Efficiency of extraction reduction ( % ) = ( C s − C o C o ) × 100   at   t   for   i
where i is the type of heavy metal (or As), t is the stabilization time (hour), Co is the concentration (mg/L) of the i component in the extracted solution (mg/L) after stabilization without AMDS (0 wt.%), and Cs is the concentration (mg/L) of the i component in the extracted solution after reaction with the AMDS. Batch experiments were repeated three times and their average value was used as the final result for the experiment.

3.4. Column Experiments for the AMDS Stabilization in Non-Equilibrium Conditions

Because the previous laboratory work in the study was limited to batch experiments in conditions of equilibrium, it was necessary to confirm the feasibility for the two types of AMDS to stabilize As and heavy metals in non-equilibrium conditions by working with them in the field. Continuous column experiments were performed to investigate the As extraction reduction efficiency of the two types AMDSs. A glass column (5 cm in diameter and 30 cm in length) was used in the column experiment, and the schematic of the column experiment is shown in Figure 3. The glass column was packed with only As-contaminated soil (soil C) and the other column was packed with As-contaminated soil (soil C) + 5 wt% (or 7%) of AMDS stabilizers. The distilled water titrated at pH 6.0 (simulating acid rain) was flushed through the column by an up-flow system at 0.2 mL/min for a 16 pore volume (about 48 months by considering the average annual precipitation). The effluent water was collected from the top of the column and 5 mL of the effluent water was sampled at certain time intervals (1 pore volume of 144 mL representing 3 months of rainfall), and its As concentration was analyzed by ICP-MS. The As extraction reduction efficiency for the two types of AMDS was quantitatively calculated using Equation (1) at each sampling time and by plotting them vs. flushing time (pore volume), and their As stabilization effects were compared.

4. Results and Discussion

4.1. Evaluation of the Characteristics of the Soils and AMDS as the Stabilizer

To identify the degree of soil contamination at four sites and the main pollutants at each site, concentrations of heavy metals in each soil sample were measured, and the results were compared with the Korean Soil Pollution Warning Limit (KSPWL) [48] (Table 2). Soil A was contaminated by Cd, Pb and Zn, resulting from the illegal dumping of metal wastes over a long time. The main pollution scenario of B site and C site was related to the mining activity. The soil sample at B site was mainly contaminated by As, which had a concentration of more than 80 times that of the KSPWL. The soil at C site was taken from the farmland near an abandoned metal mine, which was contaminated by Cd, Pb and As because of the inflow of mine tailings. The soil sample taken at D site was the surface soil in the stream next to the uphill area of the military explosion training site, which was mainly contaminated by Zn.
The XRF/XRD analyses of the MMDS and the CMDS were performed to identify their compositional and mineralogical characteristics (Figure 4 and Table 3). From the result of XRF analysis of the MMDS, the Fe2O3 content ranked at the top of the compound list at 33.20%, followed by the CaO content at 27.80% (Table 3). Similar to the MMDS, the Fe2O3 content in the CMDS was the highest at 44.23%, followed by CaO at 33.81%. The results of XRF analyses revealed that both AMDSs were composed of a large amount of Fe- and Ca-bearing minerals. The peaks of CaCO3 (calcite) and CaSO4∙2(H2O) (gypsum) were also found in the XRD analysis of the MMDS, and the peaks of CaCO3 (aragonite and calcite) in the CMDS (Figure 4). Even though both types of AMDS included high amounts of Fe2O3 in the XRF analysis, there were no distinct peaks of specific Fe-bearing minerals in the XRD analysis. At a pH lower than pH 7, most of the Fe in solution was likely to be precipitated in the form of amorphous ferrihydrite or goethite at surface water [46]. Results of XRF/XRD analyses showed that both types of AMDS were mainly composed of amorphous silicate and oxide (hydroxide) minerals as well as calcite and gypsum.
The existence of some volatile components in the MMDS and the CMDS was identified through the weight loss peaks clearly appearing in the TG-DTA analysis. For the MMDS, in a temperature range of 90–150 °C, the mass weight of the MMDS decreased by 5.144% due to the H2O evaporation and due to pyrolysis of Fe(OH)3 and Al(OH)3. In a temperature range of 570–700 °C, the decomposition of CaCO3 into CaO and CO2 resulted in a weight reduction of 6.553% (Figure 5a). The mass of the CMDS decreased by 9.383% due to the H2O evaporation at a temperature range of 50–220 °C, and the mass weight reduced by 15.596% caused by the decomposition of CaCO3 at a temperature range of 540–720 °C. In the CMDS, a tiny weight loss of 0.199% was found at 410–425 °C, which was due to the decomposition of Ca(OH)2 into CaO and H2O (Figure 5b).
Results of the BET analyses for the AMDSs are shown in Table 4. The specific surface area of the MMDS and the CMDS was 133.4 and 124.5 m2/g, respectively, suggesting that they have enough space to fix dissolved metal ions. The pore volume and average pore size of the MMDS were 0.24 cm3/g, 7.01 nm, and 0.17 cm3/g and 5.89 nm for the CMDS. The range of the pore size of the MMDS and the CMDS was 2–50 nm, and they followed the ‘type IV’ adsorption and desorption isotherms, suggesting that they were mainly composed of mesopores [57].
The SEM images and EDS peaks for the grain surface of the MMDS and the CMDS at two different magnification ratios (10,000 and 100,000 magnitudes) are shown in Figure 6. In both types of AMDS, metal hydroxide flocs were found in the granular form due to an increasing pH from the injection of neutralizing agents. The Fe2+ was oxidized during the semi-active treatment process, precipitating as the amorphous trivalent hydroxide form in the MMDS (Figure 6a,b). Granular particles in the MMDS can grow according to both the oxidation on the existing grain surface and the surface precipitation [58]. In SEM analysis for the CMDS, distinct crystal forms were observed sporadically as well as granular particles, which was because the AMD was treated by the electro-purification method (Figure 6c,d) [13]. The SEM images of both types of AMDS showed many bumpy surfaces and multiple layers originating from many irregular particles, suggesting that there were a large number of void spaces to fixate heavy metals in the AMDS (Figure 6a,c). Results of the toxicity leaching test are shown in Table 5. For the MMDS, among Cd, Pb, Ni, Zn and As, only Zn was extracted by the TCLP test and the Zn concentration in extracted solution was 0.07 mg/L. For the CMDS, Ni and Zn were extracted from the MMDS during the TCLP test, and their concentrations were below 0.07 mg/L. For the SPLP test, heavy metal barely leached out from both types of AMDS. These results confirmed that the addition of the AMDS for heavy metal stabilization in soil does not contribute to the pollution of the site. In order to evaluate the efficacy of the AMDS for the As and heavy metal stabilization in soil, each type of AMDS was added to soil A and B (heavy metal contaminated soil + 3 wt% of stabilizer) and the TCLP tests for the mixture were repeated. For only soil B, which did not have the AMDS, the As concentration of leached solution was higher than the Korean waste leaching tolerance limit (1.5 mg/L). However, for soil B mixed with 3% of the stabilizer, As was hardly leached from the soil mixture after the TCLP test (<0.01 mg/L: more than 99% reduction). The Zn concentration from the mixture of both AMDSs decreased by 15%. These results confirmed that serious heavy metal leaching from soil did not occur when the AMDS was added as a stabilizer and both of the AMDSs had a great capacity as a heavy metal stabilizer from the viewpoint of toxicity leaching. The As or heavy metal-bearing mineralization (or precipitation) on the soil particle surfaces could not be observed by SEM-EDS analysis after the stabilization experiment and it may be because the initial concentration of As and heavy metal in soil is not high and it is hard to find the newly formed mineral on the AMDS particle surface with only 5% addition into soil.

4.2. Mechanisms of as and Heavy Metal Stabilization by AMDS in Soil

The analyses of the AMDSs identified that they were mostly composed of amorphous micrometer- and sub-micrometer-sized iron(oxy)hydroxide particles and calcite. Because AMDS has a large surface area and numerous functional groups, it is suitable for use in removing dissolved heavy metals in solution through the adsorption and the co-precipitation at high pH conditions. According to previous studies [46,59,60], the hydrolysis reactions of Zn2+, Pb2+ and Cd2+ occurred in solution at a high pH level, resulting in the formation of MOH+ and hydroxides such as M(OH)2 (M: metal ions) (Equations (2) and (3)).
M2+ + H2O ↔ MOH+ + H+
MOH+ + H2O ↔ M(OH)2 + H+
The point of zero charge (PZC) of the AMDS ranged from 4.5 to 5.0 of pH, suggesting that when the pH increases up to higher than 5.0, the surface of the AMDS becomes negatively charged. The surfaces of metal(oxy)hydroxide in the AMDS are also capable of adsorbing or dissociating H+ to acquire a net surface charge according to the pH in solution (Equations (4) and (5)).
R≡MOH + H+ ↔ R≡MOH2+
R≡MOH ↔ R≡MO− + H+
As the pH increases, the functional group, MO− (Equation (5)), becomes dominant and increases the removal of heavy metals by the electrical attraction. These charged surfaces not only affect the electrostatic attraction to other heavy metal ions but also create sites for further reactions. In addition, because of the high content of Ca2+, CO32− and HCO3− in the AMDS, heavy metals are likely to be precipitated as carbonate through carbonation [61] by the following Equations (6)–(8).
3Pb2+ + 2CO32− + 2H2O → Pb2(CO3)2(OH)2 + 2H+
6Zn2+ + 2HCO3− + 6OH- → Zn6(CO3)2(OH)6 + 2H+
Cd2+ + HCO3− → CdCO3 + H+
The pH of soil A before the addition of the AMDS was 8.01 and the pH after the experiment was between 7.99 and 8.50. When the pH is between 7.22 and 9.69, the Pb exists mainly in the form of PbOH+ and Pb(OH)2 in an aqueous system. Because the PZC of the AMDS is 4.5–5.0, the surface of the AMDS is negatively charged at the pH range of soil A and thus the Pb2+ attraction on the negatively charged AMDS surfaces actively occurs. In addition, the co-precipitation of the Pb can be induced by reacting with CO32− in solution in the form of Pb2(CO3)2(OH)2. From these mechanisms, the Pb stabilization of the AMDS in soil can be activated. The Zn exists mainly in the form of ZnOH+ and Zn(OH)2 when the pH is between 7.7 and 9.1, and similar to the Pb case, the Zn5(CO3)2(OH)6 can be mainly co-precipitated with CO32− and/or OH− as well as the Zn adsorption on the AMDS surface, which has a negative charge. The pH of soil D before the experiment was 6.15, and the pH after the addition of AMDS ranged between 6.4 and 8.37. Like soil A, the Zn2+ in soil D was stabilized by both the adsorption on the negatively charged AMDS surface and/or the co-precipitation as the form of Zn5(CO3)2(OH)6 during the CaCO3 forming. Soil D contains lower concentrations of heavy metals than soil A, and thus it has the high extraction reduction efficiency of Zn due to less adsorption competition among heavy metal ions, resulting in the higher Zn stabilization efficiency of the AMDS added to soil D rather than to soil A.
The pHs of soil B and soil C were 7.26 and 7.91 before the experiment, respectively, and their pHs after the addition of the AMDS remained between 7.3 and 8.4. At this pH range, the surface of the AMDS added to soil was considered to remain negatively charged. Unlike other heavy metals, the As is likely to exist as an oxidized substance such as the arsenate (H2AsO4− or HAsO42−: a valency of 5) in surface water, and the electrical repulsive force may occur between the arsenate and the AMDS surface. Therefore, it would be difficult for the As in the form of arsenate to be adsorbed on the AMDS surface by only the electrostatic attraction when the initial pH of the soil was higher than 7.2.
Although the As-oxyanion is not likely to be directly adsorbed onto the negatively charged surface of the AMDS, the adsorption of the As is still practicable if other cations connect the As-oxyanions and the negatively charged AMDS surface as the form of a cation bridge. According to a previous study, Lin et al. (2004) observed that divalent or trivalent cations such as Ca2+, Mg2+, Fe3+, Al3+, and Mn2+ can contribute to linking between the As-oxyanions and negatively charged surfaces [62]. Fischel et al. (2015) also mentioned that as the pH of the system increases, the surface of the material becomes more negative, and it promotes adsorption of Mn2+ and Mn3+ on the solid surface, which acts as the cation bridge for the As-oxyanions [63]. Antelo et al. (2015) demonstrated that the Ca2+ obviously promotes the adsorption of arsenate on Fe-oxides such as ferrihydrite through electrostatic interaction when the pH value is above 8, and Chen et al. (2022) also observed this phenomenon under pH 4–9 conditions [64,65]. These studies suggested that cations such as Ca2+ and Mg2+ reduce this electrostatic repulsion and thus promote the adsorption of arsenate even on a negatively charged surface. Taking account of the results of these previous studies and the Ca2+ content of the AMDS, it was expected that the cation bridge effect could occur on the negatively charged surface of the AMDS, and the As fixation efficiency could increase to higher than a 7 pH value. It has also been reported that the Ca2+ enhances the As fixation via the calcium arsenate precipitation under high pH conditions [66,67,68,69,70,71] and co-precipitation during the formation of Ca(OH)2 and CaCO3, which occurred at a high Ca2+ concentration [31,66,72,73,74]. The As fixation on the AMDS also occurs due to ligand exchange and formation of new complexes between Fe-oxides and As-oxyanions [75]. When the concentration of the As in solution is low, the As mainly forms a monodentate complex on the surface of iron oxide [76], and the As stabilization of the AMDS can also occur through various Fe-As oxide complex formation processes as well as through the monodentate complex because of its high Fe content in AMDS (Table 2).

4.3. Batch Extraction Experiment for the Stabilization Efficiency of Heavy Metals and as by AMDS

Figure 7 shows the efficiency of extraction reduction (the stabilization efficiency) of the MMDS and the CMDS for Pb and Zn in soil A, which was calculated using equation (1). The results of the batch extraction experiments revealed that the Pb stabilization efficiency of the AMDS was higher than 90% at all mixed ratios of the stabilizer with soil A (Figure 7a). For the CMDS, the Pb stabilization efficiency remained higher than 80% when more than 3% of the CMDS was added (Figure 7b). Figure 7c showed that the Zn stabilization efficiency of both AMDSs was slightly lower than Pb stabilization. Park (2015) reported that when the concentration of each heavy metal in the soil is low, heavy metals show a high adsorption capacity on the adsorbent, but when the concentration of the heavy metal in soil is high, the adsorption capacity of some heavy metals to the adsorbent decreases in general because of the adsorption competition [77]. It has also been reported that this selectivity of adsorption of heavy metals on the adsorbent is related to the hydrolysis constant of the heavy metal component [78]. Therefore, it was supposed that since Pb has higher selectivity than Zn, the Zn stabilization efficiency of the AMDS is lower than Pb in soil A.
Figure 8a,b show the As stabilization efficiency of the AMDS and the CMDS in soil B and soil C. Regardless of soil type and As concentration in soil, the As stabilization efficiency of the two types of the AMDS was very high (>90%) when more than 3% of the AMDS was added to the soil. Figure 8c shows the Zn stabilization efficiency of Zn in soil D was higher than 78% for both types of AMDS, which was higher than in soil A, and it resulted from the low content of other heavy metals in soil D and from their lower adsorption competition with the AMDS compared to soil A.

4.4. Continuous Column Experiments for the AMDS Stabilization at Non-Equilibrium Condition

The mobility and stabilization efficiency of the AMDS for As and heavy metals in soil C at non-equilibrium were investigated through continuous column experiments. Figure 9 shows the experimental results of continuous column experiments with 5% and 7% of MMDS addition. In the column experiments using only As-contaminated soil C (control), the As concentration of the column effluent was 0.05~0.38 mg/L. When 5% of MMDS was added to the soil, the As concentration of the extracted solutions from the column remained below 0.01 mg/L for 16 pore volume flushing, and the pH of the effluent was 7.9~8.4 (Figure 9a,b). When compared with that of the control column, the As stabilization efficiency of both of the 5% and the 7% MMDS column was about 50~60% at the early stage of the flushing, and remained at over 60% (mostly > 70%) after 6 pore volume flushing (Figure 9c).

5. Conclusions

Even previous research has studied the possible uses for the AMDS as the adsorbent to remove heavy metals in solution, studies on their feasibility as a soil stabilizer for As and heavy metals have only just begun. Information on the mechanisms by which As and heavy metal stabilization occurs with AMDS in soil has also been very limited. This study supported the quantitative evaluation of the stabilization effect of AMDS on soil with various pollution scenarios and presented information about the major stabilization mechanisms by which As and heavy metals in soil are stabilized by AMDS. Results of various physico-chemical analyses and the toxicity leaching tests confirmed that MMDS and CMDS have a great As and heavy metal fixation capacity as soil stabilizers.
For three different types of soils contaminated with heavy metals, the Pb and Zn stabilization efficiency of both types of the AMDS was higher than 90% and 70%, respectively, due to the adsorption by the electrostatic attraction on the AMDS surfaces and the co-precipitation with carbonate and calcium hydroxide. For As-contaminated soils, the As stabilization efficiency of the AMDS was very high (>95%) with the addition of more than 3% of AMD to the soil. There can be a great difference in the stabilization efficiency of AMDSs due to different soil properties and the adequate stabilizer should be selected for the specific site. This study was focused on presenting the very high stabilization efficiency of the AMDS for various soils having different pollution scenarios and different physicochemical properties. The overall stabilization efficiency of two AMDSs for heavy metals and As were higher than 75% and 85% (mostly > 90%) with the addition of more than 3% of AMD, respectively, regardless of soil and metal type.
The main As stabilization mechanisms of AMDS were quite different from those of heavy metals and they were mainly the formation of a mono- or bi-dentate complex of arsenate with iron-oxides and/or the cation bridge formation between the arsenate and the negatively charged AMDS surface rather than only the adsorption by the electrostatic attraction.
We evaluated the economic efficiency of the stabilizer based on the raw material price, production cost, and stabilization efficiency. The unit price of the stabilizer was calculated as ‘= (purchase value of raw materials + production/modification cost)/stabilization efficiency’. This assessment confirms that the field application possibility for actual contaminated sites can be also evaluated based on the economic assessment [79]. For both the MMDS and the CMDS, the stabilization efficiency was high in both As and heavy metals, and their unit purchasing prices were in the range of 0.1–0.5 US dollar/kg, which confirms that the AMDS is superior to other commercialized stabilizers from the viewpoint of both stabilization efficiency and cost. The results from this study indicate that the AMDS could be used as an effective As and/or heavy metal stabilizer for soil with various pollution scenarios.

Author Contributions

M.L. and S.W. conceived and designed the experiments; H.T., K.K. and S.K. performed the experiments; H.T. and K.K. performed the data analyses; M.L. and S.W. contributed materials and data interpretation; M.L. and H.T. wrote the paper. All authors have read and agreed to the published version of the manuscript.

Funding

This research was supported by the Institute for Korea Spent Nuclear Fuel (IKSNF) and National Research Foundation of Korea (NRF) grant funded by the Korea government (Ministry of Science and ICT, MSIT) (2021M2E1A1085202) and by the grant (2019002470003) from Korea Ministry of Environment as “The SEM (Subsurface Environmental Management) project”.

Data Availability Statement

The data presented in this study are available on request from the corresponding author.

Conflicts of Interest

The authors declare no conflict of interest.

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Figure 1. The sampling locations of contaminated soils AMDSs for the study.
Figure 1. The sampling locations of contaminated soils AMDSs for the study.
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Figure 2. The AMD treatment process at the Ilkwang mine (a) and at the Hambaek mine (b).
Figure 2. The AMD treatment process at the Ilkwang mine (a) and at the Hambaek mine (b).
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Figure 3. Schematic of the continuous column experiments.
Figure 3. Schematic of the continuous column experiments.
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Figure 4. XRD peaks of MMDS (a) and CMDS (b).
Figure 4. XRD peaks of MMDS (a) and CMDS (b).
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Figure 5. The mass weight change of MMDS (a) and CMDS (b) with temperature increase in the TG-DTA analysis.
Figure 5. The mass weight change of MMDS (a) and CMDS (b) with temperature increase in the TG-DTA analysis.
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Figure 6. SEM images and EDS peaks of MMDS (a,b,e) and CMDS (c,d,f).
Figure 6. SEM images and EDS peaks of MMDS (a,b,e) and CMDS (c,d,f).
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Figure 7. Pb stabilization efficiency of MMDS (a) and CMDS (b) and Zn stabilization efficiency (c) in batch extraction experiments with soil A.
Figure 7. Pb stabilization efficiency of MMDS (a) and CMDS (b) and Zn stabilization efficiency (c) in batch extraction experiments with soil A.
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Figure 8. As the stabilization efficiency of AMDSs for soil B (a) and soil C (b) and Zn stabilization efficiency (c) for soil D.
Figure 8. As the stabilization efficiency of AMDSs for soil B (a) and soil C (b) and Zn stabilization efficiency (c) for soil D.
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Figure 9. As concentration (a) and the pH (b) of the effluent, and the As stabilization efficiency of MMDS (c) in the column experiments.
Figure 9. As concentration (a) and the pH (b) of the effluent, and the As stabilization efficiency of MMDS (c) in the column experiments.
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Table
Table 1. Water quality of the AMD originating from the Ilkwang mine (MMDS sampling site).
Table 1. Water quality of the AMD originating from the Ilkwang mine (MMDS sampling site).
pHConcentration of Metal and Metalloid (mg/L)
FeMnAsCdCuPbZn
AMD draining into the treatment system2.68218.98.90.650.2218.30.118.7
Korean discharge tolerance limit [50]5.8–8.62.02.00.050.021.00.11.0
Table
Table 2. Properties and heavy metal (and metalloid) concentration of 4 soil samples.
Table 2. Properties and heavy metal (and metalloid) concentration of 4 soil samples.
Soil TypeSoil
Texture (Clay
Ratio: %)		Dry Bulk Density (g/cm3)CEC (cmol/kg)Organic Content (%)pHHeavy Metal and Metalloid
Concentrations (mg/kg)
CdPbNiZnAs
ALoamy sand (11%)1.315.91.27.9536.61606.1119.54796.975.0
BLoam (22%)1.4118.72.76.233.539.611.0133.22225.3
CLoamy sand (16%)1.3614.23.17.8257.9337.812.6245.9136.3
DSand (4%)1.294.90.67.323.476.96.6704.06.3
KSPWL-----1040020060050
Bold number: >KSPWL; ‘-‘: No tolerance limit.
Table
Table 3. Principal components of AMDSs from the XRF analysis.
Table 3. Principal components of AMDSs from the XRF analysis.
Content of Component (wt. %)
Al2O3CaOCuOFe2O3K2OMgOMnOSiO2SO3SrOZnO
MMDS8.927.82.733.2-7.217.39.7-2.2
CMDS933.8-44.20.141.27.20.40.1-
-: <0.1.
Table
Table 4. Physical properties of AMDS.
Table 4. Physical properties of AMDS.
TypeSurface Area (m2/g)Pore Volume (cm3/g)Pore Diameter (nm)
MMDS133.40.247.0
CMDS124.50.175.9
Table
Table 5. TCLP and SPLP tests for AMDS.
Table 5. TCLP and SPLP tests for AMDS.
MaterialsLeaching TestHeavy Metal Concentrations (mg/L)
CdPbNiZnAs
MMDSTCLP---0.07-
SPLP-----
CMDSTCLP--0.070.01-
SPLP-----
Only soil ATCLP0.250.870.1873.320.03
Soil A + MMDS (3%)TCLP0.380.240.1863.84-
Soil A + CMDS (3%)TCLP0.300.510.2459.82-
Only soil BTCLP0.03-0.041.971.55
Soil B + MMDS (3%)TCLP0.15-0.0717.40-
Soil B + CMDS (3%)TCLP0.010.020.111.95-
Waste leaching test limit 0.33--1.5
‘-‘ represent <0.01.
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Tak, H.; Kim, S.; Kim, K.; Wang, S.; Lee, M. Stabilization of As and Heavy Metal-Contaminated Soils by Two Mine Drainage-Treated Sludges. Minerals 2023, 13, 148. https://doi.org/10.3390/min13020148

AMA Style

Tak H, Kim S, Kim K, Wang S, Lee M. Stabilization of As and Heavy Metal-Contaminated Soils by Two Mine Drainage-Treated Sludges. Minerals. 2023; 13(2):148. https://doi.org/10.3390/min13020148

Chicago/Turabian Style

Tak, Hyunji, Seonhee Kim, Kyeongtae Kim, Sookyun Wang, and Minhee Lee. 2023. "Stabilization of As and Heavy Metal-Contaminated Soils by Two Mine Drainage-Treated Sludges" Minerals 13, no. 2: 148. https://doi.org/10.3390/min13020148

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