Arsenic immobilization in water and soil using acid mine drainage sludge

Applied Geochemistry 35 (2013) 1–6

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Arsenic immobilization in water and soil using acid mine drainage sludge Myoung-Soo Ko a, Ju-Yong Kim a, Jin-Soo Lee b, Ju-In Ko b, Kyoung-Woong Kim a,⇑ a b

School of Environmental Science and Engineering, Gwangju Institute of Science and Technology (GIST), 261 Oryong-dong, Buk-gu, Gwangju 500-712, Republic of Korea Technology Research Center, Mine Reclamation Corporation (MIRECO), 30 Chungjin-dong, Jongno-gu, Seoul 110-727, Republic of Korea

a r t i c l e

i n f o

Article history: Received 4 May 2012 Accepted 6 May 2013 Available online 30 May 2013 Editorial handling by A. Mukherjee

a b s t r a c t Adsorption onto Fe-containing minerals is a well-known remediation method for As-contaminated water and soil. In this study, the use of acid mine drainage sludge (AMDS) to adsorb As was investigated. AMDS is composed of amorphous particles and so has a large surface area (251.2 m2 g1). Here, adsorption of both arsenite and arsenate was found to be almost 100%, under various initial AMDS dosages, with the arsenate adsorption rate being faster. The optimum pH for As adsorption onto AMDS was pH 7.0 and the maximum adsorption capacities for arsenite and arsenate were 58.5 mg g1 and 19.7 mg g1 AMDS, respectively. In addition, experiments revealed that AMDS dosages decreased As release from contaminated soil. Therefore, the AMDS used in this study was confirmed to be a suitable candidate for immobilizing both arsenite and arsenate in contaminated soils. Ó 2013 Elsevier Ltd. All rights reserved.

1. Introduction Arsenic is an ubiquitous trace element in the environment (Juillot et al., 1999; Matschullat, 2000) with high concentrations in some ore-forming minerals. Consequently, strong enrichment of As can occur in the environment as a result of mining and processing. Arsenic is extremely toxic (Maity et al., 2005; Elizalde-Gonzalez et al., 2001; Xu et al., 2002; Hanh et al., 2010; Lee and Yi, 2012) and long term uptake of As-enriched drinking water has been shown to cause skin lesions and gastrointestinal and cardiovascular problems, in addition it has been implicated in cancers of internal organs (Soner-Altundog˘an et al., 2000). Arsenite (As3+) and arsenate (As5+) are the common species in the environment with arsenite being 25–60 times more toxic than arsenate (Conner, 1990; Corwin et al., 1999; Pantsar-Kallio and Manninen, 1997; Moon et al., 2008). Mining and smelting processes are the main sources of As contamination in S. Korea, with the waste from abandoned Au–Ag mines being particularly significant (Chon et al., 2011). To ameliorate mine wastewater from abandoned mines, physicochemical, electrokinetic and natural attenuation processes have been employed. While these cleaning techniques have achieved high removal efficiencies of As from mine wastewater, acid mine drainage sludge (AMDS) is generated by the cleaning system. Adsorption onto Fe materials is a known to be an effective and cheap method to remediate As-rich waters, compared to ion exchange, membrane separation, bio-reduction and electrolysis ⇑ Corresponding author. Tel.: +82 62 715 2821; fax: +82 62 715 2434. E-mail address: [email protected] (K.-W. Kim). 0883-2927/$ - see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.apgeochem.2013.05.008

(Moon et al., 2008). Numerous studies have been carried out on the adsorption of As using Fe oxides (Manning et al., 1998; Swedlund and Webster, 1999; Goldberg and Johnston, 2001; Zhang et al., 2004), natural materials (Lenoble et al., 2005; Chakraborty et al., 2007; Giménez et al., 2007), synthetic materials (Chutia et al., 2009), and by-products (Genç-Fuhrman et al., 2004; Wang et al., 2008). In addition, Fe-containing materials have also been applied for As immobilization in soil. For example, Liversey and Huang (1981) and Goldberg (2002) reported that Fe-containing materials remediated an As-contaminated site as a result of adsorption processes. The objective of this study is to quantify and characterise As adsorption onto AMDS under various conditions, including AMDS dose, solution pH, reaction time and As concentration. Moreover, the feasibility of using AMDS as an As immobilization material is determined from soil experiments. 2. Materials and methods 2.1. Acid mine drainage sludge collection The AMDS used in this study was collected from Hambeak mine in S. Korea, where it had been generated from an acid mine drainage treatment system. An electrokinetic technique is applied in this treatment system to remediate the acid mine drainage with approximately 6300 tonnes of acid mine drainage being treated daily. As a result 173 tonnes of AMDS is produced annually. The AMDS samples obtained here were air dried, before being separated with respect to particle size to determine physicochemical properties. AMDS particles, which were passed through a

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M.-S. Ko et al. / Applied Geochemistry 35 (2013) 1–6

10-mesh sieve, were subject to pH, surface area (Brunauer Emmet Teller; BET), and point of zero charge (PZC) analysis; in addition, the <80-mesh AMDS particles were subject to analysis for As and X-ray diffraction (XRD). 2.2. Arsenic adsorption onto acid mine drainage sludge To determine the solid/liquid ratio for As adsorption onto AMDS, 0.1 g, 0.25 g, 0.5 g and 1.0 g of AMDS were reacted with 50 mL of an arsenite and arsenate solution, respectively. The reactions were carried out at room temperature for 24 h under agitation at 40 rpm. After 24 h, aliquots were collected and passed through a 0.45 lm filter and then stored at 4 °C prior to analysis. Arsenite and arsenate concentrations in aliquots were measured using graphite furnace atomic adsorption spectrometry (GFAAS). Differences between initial and final concentrations were subsequently used to determine the amount of As removed by various amounts of AMDS. All experiments were carried out in duplicate. A Preliminary experiment was conducted to determine the solid/liquid ratio for the adsorption process. For the pH influence experiment, the arsenite and arsenate solution was first mixed with 0.1 g AMDS. The reacted As solutions were adjusted to pH 3.0, 5.0, 7.0 and 9.0 using 0.1 M HCl or NaOH, as needed. The pH was maintained throughout the 24 h reaction period of this experiment. After 24 h, aliquots were collected and passed through a 0.45 lm filter to prepare samples for analysis and were stored at 4 °C prior to analysis. The analysis procedure followed that of the previous experiment and, again, the experiments were carried out in duplicate. For the As adsorption kinetics experiments with AMDS, various reaction times were applied to the adsorption process: 0 min, 2 min, 3 min, 6 min, 10 min, 20 min, 45 min, 90 min, 180 min, 360 min, 720 min, 1080 min and 1440 min, at initial arsenite and arsenate concentrations of 10.6 mg L1. For the kinetics experiment, 0.1 g of AMDS was mixed with 50 mL of the As solution, at pH 7.0 and room temperature under 40 rpm. Aliquots were collected after each reaction time from each set, and the preparation and analysis process were the same as in the previous experiments, and were conducted in duplicate. Adsorption isotherms were obtained to estimate the degree of adsorption the and maximum As concentration adsorbed onto the AMDS. Experiments were carried out with solutions containing various arsenite and arsenate concentrations: spiked at 10 mg L1, 20 mg L1, 100 mg L1, 200 mg L1, 600 mg L1 and 1000 mg L1. For the isotherms experimental setup, 0.1 g of AMDS was weighed, and 50 mL of spiked arsenite and arsenate solutions was then added. The series of experiments were prepared and agitated at 40 rpm at room temperature for 24 h, and the pH of each set was maintained at 7.0 using 0.1 M HCl or NaOH. The aliquots were filtered through a 0.45 lm filter after the 24 h reaction and stored at 4 °C. Arsenite and arsenate in the aliquots were measured by GFAAS and the experiments were carried out in duplicate. 2.3. Characterisation of arsenic contaminated soil Arsenic-contaminated soil samples were collected from a paddy area located in Chungyang, S. Korea, located near the abandoned Gubong Au–Ag mine (Ahn et al., 1999). Surface soils (0–20 cm depth) were sampled for this study. The summer in S. Korea has a heavy rainy season and mine waste from the Gubong mine flowed with the runoff to the paddy areas during the rainy season. This dispersion process generated As contamination. The soil samples were air dried and then passed through 10mesh (2 mm) and 80-mesh (0.18 mm) sieves. These samples were subsequently analysed for pH, size distribution, sequential extraction, and pseudo-total As using an aqua-regia extraction. The

sequential extraction method was modified after Keon et al. (2001), Kim et al. (2002) and Wenzel et al. (2001); the procedure employed five steps for the As fractions: extractable with 1 M MgCl2 (ionically bound soluble), extractable with 0.1 M NaH2PO4 (adsorbed), 0.2 M NHþ 4 -Oxalate extractable fraction (amorphous and poorly crystalline hydrous oxide of Fe and Al), 0.5 M Na-citrate + 1.0 M Na-bicarbonate extractable (crystalline Fe oxyhydroxide), and extractable with HCl and HNO3 (residual). In addition, the soil samples were examined for cation exchange capacity (CEC) and organic content using the loss-on-ignition (LOI) method. In brief, the CEC procedure was performed using 1 M Na-acetate and NH4-acetate, with about 5 g of soil mixed with 3 mL of Na-acetate for 5 min. The suspensions were then centrifuged at 2,000 rpm and discarded. This procedure was repeated four times and the soil was washed with 95% ethanol. In order to extract Na+, 3 mL of NH4-acetate was added; the suspensions were again centrifuged at 2000 rpm for 5 min. This step was repeated three times to determine the Na+ concentration. The CEC calculation is shown in

CEC ðmeq=100 gÞ ¼ ðNaþ  200Þ=23

ð1Þ

The LOI method has been widely used to determine the organic content of solid materials. In this method, the weight loss of the soil samples was measured before and after ignition, and the LOI level is the calculated using

LOI ð%Þ ¼ ððDW105  DW450 Þ=DW105 Þ  100

ð2Þ

where DW105 and DW450 represent the dry weight of soil samples after ignition at 105 °C for 1 h and 450 °C for 6 h, respectively.

2.4. Arsenic immobilization in contaminated soil The As-contaminated soil was amended with three different AMDS dosages: 0 wt%, 0.5 wt%, and 3 wt%. For the immobilization experiments, 20 g of soil (<2 mm) was weighed into an Erlenmeyer flask and 300 mL of DI water spiked with 0.01 M NaCl was added. The flasks were then agitated at 150 rpm and room temperature for 25 days, and aliquots were periodically taken from each experimental set. The collected aliquots were filtered through a 0.45 lm filter and stored at 4 °C prior to analysis. The pH and As concentration of the aliquots were monitored to estimate the As immobilization efficiency and all experiments were carried out in duplicate.

2.5. Analytical methods Arsenite and arsenate stock solutions were prepared in DI water using NaAsO2 and Na2HAsO47H2O in 0.01 M NaCl. The As concentrations in the solutions were measured using Graphite Furnace Atomic Absorption Spectrophotometry (GFAAS, AAnalyst 700, Perkin–Elmer, USA). In all experiments, the solution pH was measured using an Orion pH meter (720A), which was calibrated daily using three buffers: pH 4.0, 7.0, and 10.0. All aliquots were collected and filtered through 0.45 lm syringe filters (Whatman, Nylon) prior to arsenite and arsenate analysis. In addition, all chemicals used in this research were of analytical grade, and all stock solutions were prepared with DI water from a Milli-Q water system. The PZC of AMDS was determined using a Zetasizer NanoZS (Malvern, UK), and XRD patterns were measured using a Siemens D5005 diffractometer; the AMDS surface area was determined using a Micrometrics ASAP 2010 analyser.

M.-S. Ko et al. / Applied Geochemistry 35 (2013) 1–6

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3. Results and discussion 3.1. Acid mine drainage sludge properties The pH of the AMDS sample was 7.25, and the PZC was 7.60, the AMDS having a positive charge between acidic and neutral pH conditions. The surface area of the AMDS In this study was 251.2 m2 g1, which is greater than either granular ferric hydroxide (GFH) or magnetite, which have surface areas of 206 m2 g1 and 90 m2 g1, respectively (Dixit and Hering, 2003; Guan et al., 2008). The XRD pattern of the AMDS displayed an amorphous material peak shape. with small goethite peaks being similar to granular ferric oxide (GFO) (Fig. 1). The large surface area of AMDS is strongly influenced by the amorphous phase which would be expected to be heavily involved in As adsorption and immobilization. 3.2. Arsenic adsorption characteristics onto acid mine drainage sludge Solid/liquid ratio experiments were conducted to verify the minimum ratio required to remove approximately 10 mg L1 of arsenite and arsenate. The experiments used four AMDS doses (0.1 g, 0.25 g, 0.5 g, and 1.0 g), and the initial arsenite and arsenate concentrations were 10.4 mg L1 and 10.2 mg L1, respectively. After 24 h of reaction, the arsenite and arsenate in solution were removed in all experiments (Fig. 2), at removal rates of 100% and 99.7%, respectively. The removal reactions can be described by Eqs. (3) and (4) (Banerjee et al., 2008).

H3 AsO3 þ Fe-OH ! Fe-OH-H3 AsO3

ð3Þ

H2 AsO4 þ Fe-OH ! Fe-H2 AsO4 þ OH

ð4Þ

These results were anticipated because one unit gram of AMDS provides a large number of adsorption sites. Consequently, the optimum AMDS dosage was experimentally determined to be 0.1 g, and this dosage was used for subsequent pH effect, kinetic and isotherm experiments. To assess the influence of pH on arsenite and arsenate adsorption onto AMDS the solution pH was controlled at 3.0, 5.0, 7.0, or 9.0 using 0.1 M HCl and 0.1 M NaOH, as needed. The AMDS was

Fig. 1. XRD patterns of (a) AMDS and (b) GFO.

Fig. 2. Effect of AMDS dose on the removal of As(III) and As(V). Reaction conditions: AMDS = 0.1 g, 0.25 g, 0.5 g, 1.0 g, and As(III) = 10.4 mg L1, and As(V) = 10.2 mg L1 with 0.01 M NaCl, As(III) and As(V) solution = 50 mL, reaction time = 24 h. Values shown are averages of duplicate experiments.

able to remove arsenite and arsenate under various pH conditions, e.g., above 95.0% of the arsenite and arsenate were removed from the solutions under all pH conditions (Fig. 3). The maximum removal rate of arsenite was 98.8% at pH 7.0 and the minimum value was 95.8% at pH 3.0; the arsenate removal rate was over 98.9% for all pH conditions, with the maximum rate of 99.0% obtained at pH 7.0. The adsorption of As onto ferric oxide has been reported to decrease with increasing pH (Meng et al., 2001); however, AMDS was found to adsorb arsenite and arsenate in solution under various pH conditions. Based on these experiments, pH 7.0 was chosen as the optimum condition for arsenite and arsenate adsorption, and this value was used for subsequent kinetic and isotherm experiments. In order to determine the time required to reach adsorption equilibrium, 0.1 g AMDS was mixed with 50 mL of arsenite and arsenate solutions; the solution pH was controlled at 7.0 ± 0.1 during the reaction. The adsorption of arsenite and arsenate with time is shown in Fig. 4. Equilibrium was reached within 1 min for arsenate and 20 min for arsenite; Bang et al. (2005) have previously reported that the rate of arsenate removal was faster than that of arsenite under oxic conditions. After 1 min, 0.06 mg L1 of arsenate remained in solution, with the adsorption ratio reaching 99.0%. The arsenite adsorption by AMDS was also nearly 98.0%, though the reaction took longer than that for arsenate. The concentration of arsenite remaining was 2.54 mg L1 at 1 min. The As adsorption capacity, as calculated from the kinetic experiments, did not reflect

Fig. 3. Effect of pH on As(III) and As(V) removal from solution at pH 3.0, 5.0, 7.0 and 9.0. Reaction conditions: AMDS = 0.1 g, As(III) = 10.4 mg L1 and As(V) = 10.2 mg L1 with 0.01 M NaCl, As(V) and As(III) solution = 50 mL, reaction time = 24 h. Values shown are averages of duplicate experiments.

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M.-S. Ko et al. / Applied Geochemistry 35 (2013) 1–6 Table 1 Langmuir parameters for As(V) and As(III) adsorption onto AMDS at pH 7.0 ± 0.3.

As(V) As(III)

Cmax (mg g1)

Kads (L mg1)

R2

19.7 58.5

0.630 0.339

0.991 0.993

Table 2 Physical and chemical properties of As-contaminated soil. Properties

Values

Fig. 4. Removal kinetics of As(III) and As(V) from water spiked AMDS. Reaction conditions: AMDS = 0.1 g, As(III) and As(V) = 10.6 mg L1 with 0.01 M NaCl, As(III) and As(V) solution = 50 mL, reaction time = 0, 1, 3, 6, 10, 20, 45, 90, 180, 360, 720, 1080 and 1440 min. Values shown are averages of duplicate experiments.

a

1

C

¼

1

1

Cmax K ads ½A

þ

1

Cmax

ð5Þ

where C is the amount of adsorbed arsenate and arsenite per unit mass of AMDS (mg kg1), Cmax is the maximum adsorption capacity (mg g1), [A] is equilibrium concentration of adsorbate in the solution (mg L1), and Kads is the equilibrium adsorption constant

Fig. 5. Adsorption isotherms of As(III) and As(V) onto AMDS in spiked water. Reaction conditions: AMDS = 0.1 g, 50 mL of As(V) and As(III) spiked DI water containing 0.01 M NaCl, pH 7.0 ± 0.3; reaction time = 24 h. Symbols and j denoted the experimental results and the line shows the results simulated using a sigma plot.

5.96

LOI (%)

CEC (meq 100 g1)

Particle distribution Sand (%)

Silt (%)

Clay (%)

Texture

1.21

19.8

33.0

63.2

3.8

Silt loam

Table 3 Sequential extraction and total extraction of As-contaminated soil. Elements

the actual capacity of AMDS; in contrast, the actual adsorption capacity of AMDS could be estimated from the isotherm experiments. Adsorption isotherm experiments were carried out to identify the arsenite and arsenate adsorption characteristics of AMDS. The adsorption data for a wide range of adsorbate concentrations were most accurately described by adsorption isotherms, such the Langmuir or Freundlich isotherm models. Descriptions of the affinity between the adsorbate and adsorbent, bond energies, and adsorption capacities have previously been derived from applicable isotherm equilibrium models (Ijagbemi et al., 2009). Fig. 5 shows the adsorption isotherms for arsenate and arsenite onto AMDS at pH 7.0 ± 0.3. The Langmuir equation (Eq. (5)) was applied to determine the adsorption capacities (Stumm and Morgan, 1996; Kang et al., 2004).

pH

Step 1 Step 2b Step 3c Step 4d Step 5e Sum Totalf,g a b c d e f g

Concentration (mg kg1)

Ratio (%)

15.1 ± 1.04 11.8 ± 0.11 9.3 ± 0.32 75.8 ± 1.62 26.7 ± 1.39 138.7 131.5

10.9 8.5 6.7 54.6 19.3

Ionically bound soluble. Adsorbed. Amorphous and poorly crystalline hydrous oxide of Fe and Al. Crystalline Fe oxyhydroxide. Residual. Aqua-regia extraction. Pseudo-total.

(L mg1). The values of Cmax, Kads, and the regression coefficients (R2) are presented in Table 1. The maximum adsorption capacities of arsenate and arsenite by AMDS were 19.7 mg g1 and 58.5 mg g1, respectively, and the regression coefficients (R2) for each set of conditions were higher than 0.99. Based on the results of the isotherm experiment, it could thus be confirmed that the adsorption of arsenate and arsenite onto AMDS was successfully described by the Langmuir equation. 3.3. Arsenic immobilization in contaminated soil The As-contaminated soil is a silt loam and is weakly acidic (pH: 5.96) (Table 2). The aqua-regia extractable As concentration in the soil sample was 131.5 mg kg1 (Table 3); 80.6% of the As was in crystalline Fe and residual fractions, as obtained from the sequential extraction analysis. These As fractions have a strong resistance to weathering, though 19.4% of the As could be mobile or transfer to environmental media such as ground water, rivers and plants. It is posited here that the mobile fraction of As in soil could be controlled by AMDS. Arsenic immobilization experiments were conducted using AMDS mixing ratios of 0 wt%, 0.5 wt% and 3.0 wt%. It was expected that AMDS in soil would immobilize the transferable As, which is that released in step I or II in the sequential extraction. The pH variation displayed a similar level (5.0) in all experiments; therefore, the pH is not seen to be the significant factor affecting As behaviour in this experiment. The extracted As concentration in the 0 wt% AMDS experiments increased to 32.3 lg L1 at 25 days (Fig. 6). In the presence of AMDS, the extracted As concentration decreased at all ratios. The addition of 0.5 wt% AMDS reduced the amount of As released from the soil to 12.7 lg L1, while

M.-S. Ko et al. / Applied Geochemistry 35 (2013) 1–6

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Advisory Editor Professor Ron Fuge were very appreciated and helpful in revising the original manuscript.

References

Fig. 6. Arsenic concentration and pH variation in aliquots obtained from As immobilization experiments. Reaction conditions: AMDS mixing ratio = 0, 0.5, and 3 wt%, adding spiked DI water containing 0.01 M NaCl, agitated at 150 rpm and room at temperature for 25 days. Symbols h, , and d denote the pH variation and the bar indicates the As concentration.

23.5 lg L1 of As was immobilized by 3.0 wt% AMDS, compared to the soil with no added AMDS. The As concentration extracted from soil decreased when the AMDS dose was increased, thereby confirming that the As concentration in the aliquots depends on the AMDS mixing ratio. It should be noted that this result followed the original expectation that AMDS could immobilize As in soil, indicating that the application of AMDS to contaminated soil could indeed decrease As transfer from soil to ground water and vegetation. However, in order to further apply AMDS for environmental remediation, it is necessary to conduct additional research on As immobilization in soil under various conditions, including soil type, geochemical properties and weather conditions.

4. Conclusions The present study focused on the development of novel As immobilization materials. AMDS has found to have a high adsorption capacity for arsenite and arsenate, due to its large surface area. The As adsorption experiments indicated that arsenite and arsenate could effectively be removed from a solution by AMDS under various pH conditions. Specifically, the maximum removal efficiencies of arsenite and arsenate were 98.8% and 99.0%, respectively, under a pH of 7.0. Arsenate adsorption was faster than arsenite; from the kinetic experiments, the arsenate adsorption reaction took 1 min and arsenite took 20 min to reach equilibrium. The adsorption isotherms for arsenite and arsenate onto AMDS were best described by the Langmuir adsorption model; the maximum adsorption capacities of AMDS were 58.5 mg g1 for arsenite and 19.7 mg g1 for arsenate. The As adsorption experiments suggested that AMDS could prevent As dispersion from soil to other environments, with the As immobilization in soil depending on the AMDS mixing ratio. Consequently, 3.0 wt% AMDS could reduce the As released from soil by 23.5 lg L1 compared to the soil with no added AMDS. As such based on the results obtained from these As adsorption and immobilization experiments, AMDS should be considered as a novel candidate material, especially with respect to the reuse of by-products and As immobilization. Acknowledgements The authors gratefully acknowledge the Mine Reclamation Corporation (MIRECO) for the financial support to carry out this research. In addition, suggestions from anonymous reviewers and

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