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Removal of Chromium (VI) from wastewater using bentonite-supported nanoscale zero-valent iron
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Removal of Chromium (VI) from wastewater using bentonite-supported nanoscale zero-valent iron Li-na Shi a, Xin Zhang b, Zu-liang Chen a,* a b
School of Chemistry and Material Sciences, Fujian Normal University, Fuzhou 350007, Fujian Province, China School of Medicine, Shanxi University of Chinese Medicine, Xianyang 712000, Shanxi Province, China
article info
abstract
Article history:
Bentonite-supported nanoscale zero-valent iron (B-nZVI) was synthesized using liquid-
Received 8 February 2010
phase reduction. The orthogonal method was used to evaluate the factors impacting Cr(VI)
Received in revised form
removal and this showed that the initial concentration of Cr(VI), pH, temperature, and
14 September 2010
B-nZVI loading were all importance factors. Characterization with scanning electron
Accepted 18 September 2010
microscopy (SEM) validated the hypothesis that the presence of bentonite led to a decrease
Available online 1 October 2010
in aggregation of iron nanoparticles and a corresponding increase in the specific surface area
Keywords:
g, while the SSA of nZVI and bentonite was 54.04 and 6.03 m2/g, respectively. X-ray
Bentonite
diffraction (XRD) confirmed the existence of Fe0 before the reaction and the presence of Fe
Nanoscale zero-valent iron
(II), Fe(III) and Cr(III) after the reaction. Batch experiments revealed that the removal of Cr
Cr(VI)
(VI) using B-nZVI was consistent with pseudo first-order reaction kinetics. Finally, B-nZVI
Wastewater
was used to remediate electroplating wastewater with removal efficiencies for Cr, Pb and Cu
(SSA) of the iron particles. B-nZVI with a 50% bentonite mass fraction had a SSA of 39.94 m2/
> 90%. Reuse of B-nZVI after washing with ethylenediaminetetraacetic acid (EDTA) solution was possible but the capacity of B-nZVI for Cr(VI) removal decreased by approximately 70%. ª 2010 Elsevier Ltd. All rights reserved.
1.
Introduction
Chromium (VI) is an industrial contaminant in both soil and groundwater and is also a well-known human carcinogen (Katz and Salem, 1994). Due to its toxicity, Cr(VI) must be removed from wastewaters prior to discharge into aquatic environments (Ju-Nam and Lead, 2008). Conventional remediation techniques typically involve reduction of Cr(VI) to Cr(III) which precipitates as chromium hydroxide or chromium iron hydroxide at high pH, followed by disposal of the resulting dewatered sludge (Ngomsik et al., 2005). Other treatments, including phytoextraction, reverse osmosis, electrodialysis, ion exchange, membrane filtration and adsorption, have also been developed to remove metals from industrial wastewaters. While these methods are useful in removing Cr(VI)
from aqueous solution, they have some limitations and it is still necessary to develop new and effective remediation techniques (Mohan and Pittman, 2006). In recent years, nanoscale zero-valent iron (nZVI) has been used to remove various groundwater contaminants. The advantages of nZVI over zerovalent iron (ZVI) include higher reactive surface area, faster and more complete reactions, and better injectability into aquifers (Li et al., 2006). However, there are still some technical challenges associated with practical applications, such as the aggregation of nZVI particles and limitations imposed by high reactivity and low stability (Liu et al., 2007). Furthermore, the agglomeration of iron particles is often unavoidable due to the extremely high-pressure drops occurring in conventional systems, which along with its lack of durability and mechanical strength limits the application of nZVI (Cumbal et al., 2003).
* Corresponding author. Tel./fax: þ86 591 83465689. E-mail address: [email protected] (Z.-l. Chen). 0043-1354/$ e see front matter ª 2010 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2010.09.025
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Recently, technologies have been developed using porous materials as mechanical supports to enhance the dis¨ zu¨m et al., 2009). Bentonite is persibility of nZVI particles (U a traditional low-cost efficient adsorbent, which has high potential for heavy metal removal from wastewaters due to its abundance, chemical and mechanical stability, high adsorption capability and unique structural properties (Bhattacharyya and Gupta, 2008). Removal of metal ions using bentonite is based on ion exchange and adsorption mechanisms due to the materials relatively high cation exchange capacity (CEC) and specific surface area (Bhattacharyya and Gupta, 2008). In this paper, bentonite was used as a porousbased support material for synthesized nZVI. More recently, nZVI supported by zeolite (Li et al., 2007) and stabilized by chitosan (Geng et al., 2009) has been reported to increase the durability and mechanical strength of nZVI. However, only a few studies have reported using natural clays as support ¨ zu¨m et al., 2009). materials for nZVI (U In this paper the removal of Cr(VI) from an aqueous solution was investigated using B-nZVI and the objectives were: (1) synthesis of bentonite-supported nanoscale zero-valent (B-nZVI) by reduction of Fe3þ ions with NaBH4, and characterization of the produced material with SEM, XRD and BET-N2 technology; (2) evaluation of the factors impacting on Cr(VI) removal using an orthogonal method; nZVI and bentonite were used for Cr(VI) removal individually as a control, and the kinetics of Cr(VI) reduction by B-nZVI were also evaluated; and (3) remediation of electroplating wastewater including some heavy metal ions using B-nZVI and evaluation of reuse.
2.
Materials and methods
2.1.
Materials and chemicals
Bentonite was provided by Fenghong Co. Ltd, Anji, Zhejiang, China, primarily as Na-Mt montmorillonite (>90%), the chemical composition was 62.5% SiO2, 18.5% Al2O3, 1.75% Fe2O3, 4.25% MgO, 0.95% CaO, and 2.75% Na2O. The cation exchange capacity (CEC) was 75.4 meq/100g. After drying overnight at 80 C, the raw bentonite was ground and sieved through a 200 mesh screen prior to use in experiments. All the reagents were analytical grade (Shanghai Nanxiang Reagent Co., Ltd., China) and distilled water was used in all preparations. A stock solution containing potassium dichromate (K2Cr2O7) was prepared by dissolving K2Cr2O7 with deionized water and a series of solutions used during the experiment were prepared by diluting the stock to the desired concentrations.
2.2.
Synthesis of nZVI and supported nZVI
The nZVI and B-nZVI were prepared using conventional liquid-phase methods via the reduction of ferric iron by borohydride without or with bentonite as a support material (Celebi et al., 2007). Bentonite (2.00 g) was initially placed into a three-necked open flask, and a ferric solution produced by dissolving ferric chloride hexa-hydrate (9.66 g) in an ethanolwater solution (50 mL, 4:1 v/v) was added and stirred for 10 min. Subsequently, a freshly prepared NaBH4 solution
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(3.54 g of NaBH4 in 100 mL) was added drop-wise into the mixture with constant stirring for 20 min after addition. The whole process described above was performed under a N2 atmosphere with vigorous stirring to avoid the oxidization of B-nZVI. The formed suspension was filtered and the black nanoscale precipitate was washed three times with pure ethanol and dried overnight at 75 C under vacuum (Celebi ¨ zu¨m et al., 2009). The theoretical mass fraction et al., 2007; U of bentonite in synthesized B-nZVI was 50%, and nZVI was prepared under identical conditions but with bentonite omitted. The nZVI and supported nZVI samples were stored in brown, sealed bottles under dry conditions and were not acidified prior to use.
2.3.
Characterizations and measurements
Scanning electron microscopy (SEM) was performed using a Philips-FEI XL30 ESEM-TMP (Philips Electronics Co., Eindhoven, The Netherlands). Images of various materials were obtained at an operating voltage of 30 kV. The SSA of nZVI, B-nZVI, and bentonite was measured via the BET adsorption ¨ zu¨m et al., 2009) using Micromeritics’ ASAP 2020 method (U Accelerated Surface Area and Porosimetry Analyzer (Micromeritics Instrument Corp.,USA). The specific surface areas of nZVI, B-nZVI and bentonite were 54.04, 39.94 and 6.03 m2/g, respectively. X-ray diffraction (XRD) patterns of B-nZVI before and after contacting Cr(VI) were performed using a PhilipsX’Pert Pro MPD (Netherlands) with a high-power Cu- Ka radioactive source (l ¼ 0.154 nm) at 40 kV/40 mA. The concentration of total Cr in solution and the concentrations of different heavy metal ions in the wastewater were determined using a flame atomic absorbance spectrometer (VARIAN AA 240FS, USA), and the Cr(VI) concentration was determined using the 1,5-diphenylcarbazide method (Geng et al., 2009) on a 722N visible spectrophotometer (Shanghai Precision & Scientific Instrument Co., Ltd, China).
2.4.
Batch experiments
The orthogonal method was used to test the effects of various factors on the reaction, and to optimize the conditions for Cr(VI) removal using B-nZVI. The experimental design was developed with the aid of the Orthogonal Design Assistant, where the initial concentration of Cr(VI), B-nZVI loading, temperature and pH were chosen as variables. Cr(VI) solutions (25 mL) with a known mass of B-nZVI were sealed in 50 mL centrifuge tubes and mixed for 4h before being centrifuged prior to analysis of the aqueous phase for residual Cr(VI). In order to investigate the role that bentonite and Fe0 played in the B-nZVI system, nZVI, B-nZVI and bentonite were all used in batch experiments examining Cr(VI) removal from aqueous solutions at an initial concentration of 50 mg/L at 35 C and 250 r/min. As the mass ratio of Fe0:bentonite was 1:1 in the B-nZVI system, the dosages of nZVI and bentonite were both set at 1.5 g/L, which was half the B-nZVI dosage. The mixtures were filtered through 0.45 mm mixed cellulose ester (MCE) membranes prior to determining the residual concentrations of Cr(VI) after contacting for 3 h. In order to investigate the effects of the different factors mentioned in
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the orthogonal experiments further, the mixtures of Cr(VI) solution and B-nZVI were mixed in 50 mL centrifugal tubes in a rotary shaker for determined periods of time, the conditions of which were initially set at 25 mL of 50 mg/L Cr(VI) solution, 3 g/L of B-nZVI, 35 C and 250 r/min. At selected timed intervals, the suspension was filtered through 0.45 mm MCE membranes, and the concentration of Cr(VI) in the filtrate was determined. To explore the feasibility of removing heavy metal ions from wastewater, B-nZVI was used to remediate electroplating wastewater collected from an electroplating factory’s sewage outfall (Fuzhou, China). The wastewater was centrifuged at 3000 r/min for 10 min to remove any insoluble impurities, prior to determining the initial pH and concentrations of total Cu, Cr, Pb and Zn. A batch of 50 mL bottles containing wastewater (10 mL) and B-nZVI (0.10 g) were mixed on a rotary shaker at 35 C and 250 r/min for 4 h. Then the mixtures were centrifuged at 3000 r/min for 10 min and the upper aliquot collected to determine pH and the concentration of each heavy metal ion. The potential to reuse B-nZVI for removing Cr(VI) from aqueous solution was also evaluated. B-nZVI (0.1 g) was added to 50 mg/L Cr(VI) solution (25 mL) and the mixture was shaken on a rotary shaker (35 C and 250 r/min). After 3 h, centrifugations at 3000 r/min were performed for 10 min to obtain solid-liquid separation. The supernatant was decanted carefully and used to determine the concentration of Cr(VI) remaining in solution while the used B-nZVI was mixed with different concentrations of EDTA. The residual B-nZVI-EDTA solution was washed with distilled water three times and shaken for another 3 h under identical conditions. The B-nZVI treated with 50 mg/L and 10 mg/L of EDTA was used to remove Cr(VI) for 3 times in succession to test the efficacy of reuse. In order to ascertain the accuracy, reliability and reproducibility of the data, orthogonal experiments were conducted in quadruplicate (n ¼ 4) and other batch experiments were carried out in triplicate (n ¼ 3) to minimize any experimental errors. The average values of the parallel measurements were used in all analysis and together with the standard deviations of these means were listed in Tables 1 and 3.
3.
Results and discussion
3.1.
Characterization
Table 1 e Orthogonal experimental design and the results obtained from the full 24 factorial experiment matrix. Removal T Cr(VI)ina pHina B-nZVIina Cr(VI)res (mg/L) (mg/L) efficiency (%) ( C) (mg/L) 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16
25 25 25 25 30 30 30 30 35 35 35 35 40 40 40 40
20 50 70 100 20 50 70 100 20 50 70 100 20 50 70 100
2.0 5.0 8.0 10 5.0 2.0 10.0 8.0 8.0 10.0 2.0 5.0 10.0 8.0 5.0 2.0
2 3 4 5 4 5 2 3 5 4 3 2 3 2 5 4
0.11a 29 1 45.0 0.6 73 2 0a 0.06a 55.3 0.9 75 2 0.04a 31.1 0.7 0.04a 71 2 0.10a 6.75a 32.0 0.9 0.24a
99.5 42.0 35.7 26.6 100 99.9 21.0 25.3 99.8 55.6 99.9 29.4 99.5 86.5 54.3 99.8
ina - initial; res - residual; a means the standard deviations are too low to be listed.
presence of bentonite (Fig. 1). The synthesized nZVI without bentonite as a support material showed that nZVI particles were roughly globular and aggregated into a chain-like conformation (Fig. 1a). The diameters of the nanoscale zerovalent iron particles were in the range of 20e90 nm when bentonite was introduced as a support material. Compared with Fig. 1a, the aggregation of nZVI particles seemed to decrease and their dispersity increase in Fig. 1b, where the mass fraction of bentonite was 50%. A similar conclusion has been drawn using kaolin as a support material to synthesize kaolin supported nZVI, which was used to remove Cu(II) and Co ¨ zu¨m et al., 2009). As indicated in (II) from an aqueous solution (U Fig. 1c, the sizes of iron nanoparticles increase prominently after reacting with Cr(VI). This phenomenon could be attributed to the co-precipitation of Cr(III) and Fe(III) on the surface of the nanoparticles (Ponder et al., 2000; Manning et al., 2007), which occurs due to a redox reaction between Cr(VI) and Fe0 (Ponder et al., 2000; Manning et al., 2007). The XRD patterns of synthesized materials were compared with the XRD patterns obtained from standard materials, to identify the apparent peaks attributable to different iron and chromium compounds. The XRD patterns of B-nZVI before reaction (Fig. 2a) with Cr(VI) showed an apparent peak of Fe0 (2q z 44.90), which weakened significantly after the reaction
The SEM images of nZVI and B-nZVI showed the morphology and nanoparticle distribution of nZVI in the absence or
Table 2 e Range analysis and variance analysis of the orthogonal test. Factors
Temperature ( C) Cr(VI)ina (mg/L) pHina B-nZVIina (mg/L)
Range Analysis
Variance Analysis
k1
k2
k3
k4
Ranges
SSE
DOF
F-value
F critical values
50.9 99.7 99.7 59.1
61.6 71.0 56.4 66.7
71.2 52.7 61.8 72.8
85.0 45.3 50.7 70.1
34.1 54.4 49.1 13.7
2518 7039 5912 422
3 3 3 3
0.63 1.77 1.49 0.11
3.49 3.49 3.49 3.49
ina - initial; SSE - the Square Sum of Errors; DOF - Degree of Freedom.
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optimal conditions for chromium removal were 40 C, 20 mg/L of initial Cr(VI) concentration, 4 g/L B-nZVI loading and pH 2.0.
Table 3 e Remediation of actual electroplating wastewater by B-nZVI. Wastewater
Total Cr
Pb2þ
Cu2þ
c0 (mg/L) 73 2 13.1 0.6 33 1 a 2.4 0.1 c’ (mg/L) after reaction a Removal amount 7.3 1.3 3.0 (mg/g B-nZVI) Removal percentage (%) 100 100 92.7
Zn2þ
pH
284 4 1.9 115 3 4.5 16.8 e 59.4
e
a means the concentration of the heavy metal was under the limit of detection.
¨ zu¨m et al., 2009). The XRD patterns of B-nZVI after (Fig. 2b) (U reaction (Fig. 2b) indicated the presence of g-Fe2O3 (2q ¼ 35.68), Fe3O4 (2q ¼ 35.45) and Cr2FeO4 (2q ¼ 35.50), which were not detected in the sample before reaction (Chen et al., 2008). The appearance of Fe(II), Fe(III) and Cr(III) in B-nZVI after reaction demonstrated the occurrence of redox reactions between Fe0 and Cr(VI) where nZVI particles were acting as reductants, which was consistent with previous literature (Ponder et al., 2000).
3.2.
889
Orthogonal test
The designed complex conditions and the final removal efficiencies of Cr(VI) from aqueous solution by B-nZVI were listed in Table 1. Results were processed using the Orthogonal Design Assistant software (El Hajjouji et al., 2008) in Range Analysis and Variance Analysis (Table 2). The higher the range and the F-value, the more significant the factor was and the greater the influence of the factor on Cr(VI) removal Comparing the ranges and F-values in Table 2, factors influencing Cr(VI) removal were (in order of decreasing influencing): initial Cr(VI) concentration > pH > temperature > BnZVI loading. The K values from K1 to K4 represented each level of each factor, from the lowest to the highest. The higher the K value, the higher the removal efficiency, and the better the level of the factor. Take temperature for example, the highest K was K4, which represented the level 40 C, and this made 40 C the optimum temperature. The change in K values indicated that chromium removal increased with temperature and decreased as initial Cr(VI) concentration and pH rose. The
3.3.
Conditions affecting Cr(VI) removal
After contacting for 3 h under identical conditions, the removal efficiencies of Cr(VI) were 5.5, 60.0 and 100.0% respectively when bentonite, nZVI and B-nZVI were added individually. Bentonite generally has poor adsorption of Cr(VI) due to its negatively charged surface and the predominant existence of Cr(VI) as anions (Bhattacharyya and Gupta, 2008). The activity of nanoscale zero-valent iron particles was enhanced significantly when bentonite was introduced as a support material, which confirmed the role bentonite played as a dispersant and stabilizer in B-nZVI (Ponder et al., 2000). Kinetics studies of Cr(VI) reduction using B-nZVI suggested that the reactivity of nZVI particles supported on bentonite were enhanced significantly due to an increase in SSA and a decrease in aggregation. Reduction kinetics of Cr(VI) by B-nZVI were described by a pseudo first-order reaction (Ponder et al., 2000): dc v ¼ ¼ kSA as rm c dt
(1)
Where c was the concentration (mg/L) of contaminant, kSA was the specific reaction rate constant associated with the SSA of the materials (L/h m2), as was the specific surface area (m2/g), and rm was the mass concentration (g/L). For kSA, as, and rm are constant for a specific reaction, the product of the three can be expressed with one parameter kobs, which is called the observed rate constant of a pseudo first-order reaction (h1). Therefore Eq. (1) can be integrated into: ln
c ¼ kobs t c0
(2)
The kobs values under different conditions are equal to the slope of the line achieved by plotting lnðc=co Þ versus time under various conditions. In this study, the plots of lnðc=co Þ versus time produced linear plots with correlation coefficients (R2) higher than 0.9 (Fig. 3). This indicated that the rate of Cr(VI) reduction by B-nZVI fitted well the pseudo first-order model under various conditions. Additionally, the reduction of Cr(VI) by B-nZVI represented a solid-liquid inter-phase reaction, which agreed with the pseudo first-order kinetics model.
Fig. 1 e SEM images of laboratory synthesized iron particles with and without a support material. a. NZVI; b. B-nZVI before reaction with Cr(VI) solution; c. B-nZVI after reaction with Cr(VI) solution. The scale bar in the figure is 500 nm.
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3.3.1.
Fig. 2 e X-ray diffractogram of B-nZVI. a. before the reaction with Cr(VI) solution; b. after the reaction with Cr (VI) solution.
Effect of initial Cr(VI) concentration
The effect of initial Cr(VI) concentration on removal efficiency was investigated in the range 20e100 mg/L. The plot fitted the pseudo first-order model well (Fig. 3a), where the observed rate constant decreased significantly as the initial Cr(VI) concentration increased, which agreed with the orthogonal test results. The equilibrium time became longer and the final removal efficiency of Cr(VI) decreased as the initial Cr(VI) concentration increased, so that while the percentage of Cr(VI) removed within 20 min at a Cr(VI) concentration of 20 mg/L was nearly 100%, it was only 30.4% within 60 min at a Cr(VI) concentration of 100 mg/L. Generally, the slower rate and lower efficiency of Cr(VI) removal from aqueous solution were found at higher concentrations of Cr(VI). Based on the SEM analysis and previous research, Cr(VI) reduction by nZVI could be defined as a surface-mediated process (Ponder et al., 2000; Rivero-Huguet and Marshall, 2009). The more the Cr(VI) ion approached the surface of nZVI dispersed on the bentonite, the faster Fe0 was oxidized into Fe(III) and the faster the coprecipitation of Cr(III) and Fe(III) oxides/hydroxides occurred. This reduced the reactivity of nZVI and subsequently resulted
Fig. 3 e Effects of various factors on Cr(VI) removal by fitting to the pseudo first-order model. a. initial Cr(VI) concentration; b. B-nZVI loading; c. pH value; d. temperature.
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in a decrease in kobs (Geng et al., 2009; Yuan et al., 2009). On the other hand, the highest removal amount was obtained at an initial Cr(VI) concentration of 70 mg/L, which could be ascribed to the limited capacity of B-nZVI for Cr(VI) removal determined by SSA.
3.3.2.
Effect of B-nZVI loading
The initial loadings of B-nZVI in Cr(VI) solution were 1, 2, 3 and 4 g/L. The Cr(VI) removal percentage increased as the B-nZVI concentration increased (Fig. 3b). The removal percentage of Cr(VI) was 54.6% using B-nZVI at 1 g/L for 120 min, but was nearly 100% when the B-nZVI loading was over 3 g/L. Meanwhile, kobs increased as the B-nZVI loading increased. These phenomena can be attributed to the increase in the available active sites resulting from the elevation in B-nZVI loading, where the reduction of Cr(VI) occurred (Geng et al., 2009; Yuan et al., 2009). However, the concentration of Cr(VI) decreased dramatically in the initial 10 min, then slightly declined in the later reduction. A few researchers (Ponder et al., 2000; Manning et al., 2007) reported a sorption phase during the reaction which could also be supported by our SEM images, suggesting that the overall mechanism was more complicated than a simple chemical reaction.
3.3.3.
Effect of the pH value
The two dominant forms of Cr(VI) in aqueous solution were 2 HCrO 4 , between pH 1.0 to 6.0 and CrO4 above pH 6.0 (Mohan and Pittman, 2006). The dependence of the reaction rate constant on pH was investigated by adjusting the solution pH to 2.0, 4.0, 6.0 and 8.0 with either 0.1 M HCl or NaOH (Yuan et al., 2009). Except for pH 2.0, the reduction of Cr(VI) can be described using the pseudo first-order model well (Fig. 3c). A remarkable increase in the removal rate occurred at pH 2.0, where equilibrium was achieved within 1 min and the residual Cr(VI) was below the detection limit, which made kinetic fitting infeasible. The Cr(VI) removal percentage decreased significantly with increases in the initial pH, so that only 27.2% Cr(VI) was reduced at pH 8.0 in 20 min while nearly 100% Cr(VI) was removed in 1 min at pH 2.0. In addition, kobs was respectively 0.0275, 0.0163, and 0.0083/min, when the initial pH value was 4.0, 6.0 and 8.0, indicating that the reduction rate increased as pH decreased, which had also been reported in other studies (Geng et al., 2009; Yuan et al., 2009). These results demonstrated that a lower pH favored Cr(VI) reduction, since at lower pH corrosion of nZVI was accelerated and the precipitation of Cr(III) and Fe(III) hydroxides on the surface of iron was consequently not as favorable, which led to an increase in the reaction rate (Lee et al., 2003). Furthermore, the increase in Hþ concentration left the surface of bentonite less negatively charged, which reduced the electrostatic repulsion between bentonite and Cr(VI) anions. This consequently promoted the electron transfer between zero-valent iron and Cr(VI) (Yuan et al., 2009).
3.3.4.
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ambient temperature. Fig. 3d highlights the relationship between lnðc=co Þ and time, where the linearity suggested that the reduction of Cr(VI) at different temperatures in the presence of B-nZVI fitted pseudo first-order dynamics (Ponder et al., 2000; Manning et al., 2007). The kobs was 0.020, 0.023, 0.026 and 0.030/min at four temperatures (25, 30, 35 and 40 C), showing that an increase in the reaction temperature resulted in an improved reaction rate. The apparent activation energy (Ea) of Cr(VI) reduction by B-nZVI was 24.9 kJ/mol, demonstrating that it is a chemically controlled adsorption process having an Ea value higher than 21 kJ/mol (Geng et al., 2009).
3.4. B-nZVI used to remove Cr(VI) from electroplating wastewater and B-nZVI reuse The data obtained from batch experiments where B-nZVI was used to remove Cr(VI) and other metals from electroplating wastewater are presented in Table 3, which indicated that BnZVI had the capacity to remove various heavy metals and was a potential promising candidate for applications to in situ environmental remediation. After reacting 10 mL of the wastewater with 0.1 g of B-nZVI for 4 h, the residual concentration of each metal ion showed that 100% total Cr, 100% Pb (II), 92.7% Cu(II), and 59.4% Zn(II) were removed, following treatment with B-nZVI. Total Cr, Pb(II) and Cu(II) received higher removal percentages due to their higher standard reduction potentials compared to Fe(II) (fqFeðIIÞ=Fe0 ¼ 0:44V). In contrast, a lower removal percentage of Zn(II) was obtained because the standard reduction potential of Zn(II) (fqZnðIIÞ=Zn0 ¼ 0:762V) was more negative than Fe(II) (Ladd, 2004). The amount of Cr(VI) removed when using B-nZVI treated with different concentrations of EDTA after being used four times was calculated and it was shown that B-nZVI’s ability to remove Cr(VI) was dramatically reduced after being used only once with an initial Cr(VI) concentration of 50 mg/L (Fig. 4). The rapid deterioration of B-nZVI was ascribed to the inability of the redox reaction between Cr(VI) and Fe0 to proceed
Effect of temperature
To assess the effect of different temperatures, batch experiments were conducted at 25, 30, 35 and 40 C. The results showed that 82.4% Cr(VI) was removed at 40 C while only 73.4% Cr(VI) was reduced at 25 C in 60 min. Thus zero-valent iron could have a positive effect on Cr(VI) reduction even at
Fig. 4 e The variation of Cr(VI) removal amount by B-nZVI after reusing four different times. The solutions of EDTA used for treatment of B-nZVI were 50 mg/L and 10 mg/L respectively as marked in the figure.
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further since this was a chemical controlled and irreversible process. As confirmed by XRD analysis (Fig. 2), reaction products were deposited on the surface of B-nZVI in the form of oxide-hydroxide co-precipitation of Fe(II), Fe(III) and Cr(III), which consequently decreased the activity of Fe0 (Chen et al., 2008). This phenomenon confirmed that the active ingredient of B-nZVI was Fe0 which acted as a reductant, while bentonite only played a role as a dispersant and stabilizer.
4.
Conclusions
In this study, nZVI particles became more effective when bentonite was introduced as a support material due to reduction of aggregation and increased SSA. Batch experiments indicated that the removal rate increased as the temperature and B-nZVI loading (4 g/L) increased, and fell as the initial Cr(VI) concentration and pH increased, which agreed with the result obtained from the orthogonal test. Under the various operational conditions considered, reduction of Cr(VI) using B-nZVI was in accordance with a pseudo first-order model. B-nZVI was effective in removing Cr(VI) and other heavy metals, including Pb(II), Cu(II) and Zn(II) from electroplating wastewater. Since bentonite is a stable and lowcost clay mineral, B-nZVI could be an efficient and promising remediation material to remove Cr(VI) and other metals from wastewater. However, further research must be carried out to slow and control the degree of nZVI oxidation in the atmosphere and as a consequence a more effective regeneration method may emerge from such studies.
Acknowledgements This work is supported by the Fujian “Minjiang Fellowship” from Fujian Normal University (gs1). The authors also gratefully acknowledge the significant contributions of Dr Gary Owens in editing and improving the many revisions of this manuscript, his suggestions and corrections have undoubtedly significantly improved the quality of the final manuscript.
references
Bhattacharyya, K.G., Gupta, S.S., 2008. Adsorption of a few heavy metals on natural and modified kaolinite and montmorillonite: a review. Adv. Colloid Interface Sci. 140, 114e131. ¨ zu¨m, C¸., Shahwan, T., Erten, H.N., 2007. A radiotracer Celebi, O., U study of the adsorption behavior of aqueous Ba2þ ions on
nanoparticles of zero-valent iron. J. Hazard. Mater. 148, 761e767. Chen, S.S., Hsu, B.C., Hung, L.W., 2008. Chromate reduction by waste iron from electroplating wastewater using plug flow reactor. J. Hazard. Mater. 152, 1092e1097. Cumbal, L., Greenleaf, J., Leun, D., SenGupta, A.K., 2003. Polymer supported inorganic nanoparticles: characterization and environmental applications. React. Funct. Polym. 54, 167e180. El Hajjouji, H., Ait Baddi, G., Yaacoubi, A., Hamdi, H., Winterton, P., Revel, J.C., Afidi, M., 2008. Optimisation of biodegradation conditions for the treatment of olive mill wastewater. Bioresour. Technol. 99, 5505e5510. Geng, B., Jin, Z., Li, T., Qi, X., 2009. Kinetics of hexavalent chromium removal from water by chitosan-Fe0 nanoparticles. Chemosphere 75, 825e830. Ju-Nam, Y., Lead, J., 2008. Manufactured nanoparticles: an overview of their chemistry, interactions and potential environmental implications. Sci. Total Environ. 400, 396e414. Katz, S., Salem, H., 1994. The Biological and Environmental Chemistry of Chromium. VCH Publishers, New York. Ladd, M.F.C., 2004. Introduction to Physical Chemistry, third ed. Cambridge University Press, Cambridge. Lee, T., Lim, H., Lee, Y., Park, J.W., 2003. Use of waste iron metal for removal of Cr(VI) from water. Chemosphere 53, 479e485. Li, X., Elliott, D.W., Zhang, W.X., 2006. Zero-valent iron nanoparticles for abatement of environmental pollutants: materials and engineering aspects. Crit. Rev. Solid State Mater. Sci. 31, 111e122. Li, Z., Kirk Jones, H., Zhang, P., Bowman, R.S., 2007. Chromate transport through columns packed with surfactant-modified zeolite/zero valent iron pellets. Chemosphere 68, 1861e1866. Liu, Y., Phenrat, T., Lowry, G.V., 2007. Effect of TCE concentration and dissolved groundwater solutes on NZVI-promoted TCE dechlorination and H2 evolution. Environ. Sci. Technol. 41, 7881e7887. Manning, B.A., Kiser, J.R., Kwon, H., Kanels, S.R., 2007. Spectroscopic investigation of Cr (III)-and Cr (VI)-treated nanoscale zerovalent iron. Environ. Sci. Technol. 41, 86e592. Mohan, D., Pittman, C.U., 2006. Activated carbons and low cost adsorbents for remediation of tri-and hexavalent chromium from water. J. Hazard. Mater. 137, 62e811. Ngomsik, A., Bee, A., Draye, M., Cote, G., Cabuil, V., 2005. Magnetic nano-and microparticles for metal removal and environmental applications: a review. Comptes RendusChimie 8, 963e970. Ponder, S.M., Darab, J.G., Mallouk, T.E., 2000. Remediation of Cr (VI) and Pb (II) aqueous solutions using supported, nanoscale zero-valent iron. Environ. Sci. Technol. 4, 2564e2569. Rivero-Huguet, M., Marshall, W.D., 2009. Reduction of hexavalent chromium mediated by micron-and nano-scale zero-valent metallic particles. J. Environ. Monit. 11, 1072e1079. ¨ zu¨m, C¸., Shahwan, T., Ero lu, A.E., Hallam, K.R., Scott, T.B., U Lieberwirth, I., 2009. Synthesis and characterization of kaolinite-supported zero-valent iron nanoparticles and their application for the removal of aqueous Cu2þ and Co2þ ions. Appl. Clay Sci. 43, 172e181. Yuan, P., Fan, M., Yang, D., He, H., Liu, D., Yuan, A., Zhu, J., Chen, T., 2009. Montmorillonite supported magnetite nanoparticles for the removal of hexavalent chromium [Cr (VI)] from aqueous solutions. J. Hazard. Mater. 166, 821e882.
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Removal of Chromium (VI) from wastewater using bentonite-supported nanoscale zero-valent iron Li-na Shi a, Xin Zhang b, Zu-liang Chen a,* a b
School of Chemistry and Material Sciences, Fujian Normal University, Fuzhou 350007, Fujian Province, China School of Medicine, Shanxi University of Chinese Medicine, Xianyang 712000, Shanxi Province, China
article info
abstract
Article history:
Bentonite-supported nanoscale zero-valent iron (B-nZVI) was synthesized using liquid-
Received 8 February 2010
phase reduction. The orthogonal method was used to evaluate the factors impacting Cr(VI)
Received in revised form
removal and this showed that the initial concentration of Cr(VI), pH, temperature, and
14 September 2010
B-nZVI loading were all importance factors. Characterization with scanning electron
Accepted 18 September 2010
microscopy (SEM) validated the hypothesis that the presence of bentonite led to a decrease
Available online 1 October 2010
in aggregation of iron nanoparticles and a corresponding increase in the specific surface area
Keywords:
g, while the SSA of nZVI and bentonite was 54.04 and 6.03 m2/g, respectively. X-ray
Bentonite
diffraction (XRD) confirmed the existence of Fe0 before the reaction and the presence of Fe
Nanoscale zero-valent iron
(II), Fe(III) and Cr(III) after the reaction. Batch experiments revealed that the removal of Cr
Cr(VI)
(VI) using B-nZVI was consistent with pseudo first-order reaction kinetics. Finally, B-nZVI
Wastewater
was used to remediate electroplating wastewater with removal efficiencies for Cr, Pb and Cu
(SSA) of the iron particles. B-nZVI with a 50% bentonite mass fraction had a SSA of 39.94 m2/
> 90%. Reuse of B-nZVI after washing with ethylenediaminetetraacetic acid (EDTA) solution was possible but the capacity of B-nZVI for Cr(VI) removal decreased by approximately 70%. ª 2010 Elsevier Ltd. All rights reserved.
1.
Introduction
Chromium (VI) is an industrial contaminant in both soil and groundwater and is also a well-known human carcinogen (Katz and Salem, 1994). Due to its toxicity, Cr(VI) must be removed from wastewaters prior to discharge into aquatic environments (Ju-Nam and Lead, 2008). Conventional remediation techniques typically involve reduction of Cr(VI) to Cr(III) which precipitates as chromium hydroxide or chromium iron hydroxide at high pH, followed by disposal of the resulting dewatered sludge (Ngomsik et al., 2005). Other treatments, including phytoextraction, reverse osmosis, electrodialysis, ion exchange, membrane filtration and adsorption, have also been developed to remove metals from industrial wastewaters. While these methods are useful in removing Cr(VI)
from aqueous solution, they have some limitations and it is still necessary to develop new and effective remediation techniques (Mohan and Pittman, 2006). In recent years, nanoscale zero-valent iron (nZVI) has been used to remove various groundwater contaminants. The advantages of nZVI over zerovalent iron (ZVI) include higher reactive surface area, faster and more complete reactions, and better injectability into aquifers (Li et al., 2006). However, there are still some technical challenges associated with practical applications, such as the aggregation of nZVI particles and limitations imposed by high reactivity and low stability (Liu et al., 2007). Furthermore, the agglomeration of iron particles is often unavoidable due to the extremely high-pressure drops occurring in conventional systems, which along with its lack of durability and mechanical strength limits the application of nZVI (Cumbal et al., 2003).
* Corresponding author. Tel./fax: þ86 591 83465689. E-mail address: [email protected] (Z.-l. Chen). 0043-1354/$ e see front matter ª 2010 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2010.09.025
w a t e r r e s e a r c h 4 5 ( 2 0 1 1 ) 8 8 6 e8 9 2
Recently, technologies have been developed using porous materials as mechanical supports to enhance the dis¨ zu¨m et al., 2009). Bentonite is persibility of nZVI particles (U a traditional low-cost efficient adsorbent, which has high potential for heavy metal removal from wastewaters due to its abundance, chemical and mechanical stability, high adsorption capability and unique structural properties (Bhattacharyya and Gupta, 2008). Removal of metal ions using bentonite is based on ion exchange and adsorption mechanisms due to the materials relatively high cation exchange capacity (CEC) and specific surface area (Bhattacharyya and Gupta, 2008). In this paper, bentonite was used as a porousbased support material for synthesized nZVI. More recently, nZVI supported by zeolite (Li et al., 2007) and stabilized by chitosan (Geng et al., 2009) has been reported to increase the durability and mechanical strength of nZVI. However, only a few studies have reported using natural clays as support ¨ zu¨m et al., 2009). materials for nZVI (U In this paper the removal of Cr(VI) from an aqueous solution was investigated using B-nZVI and the objectives were: (1) synthesis of bentonite-supported nanoscale zero-valent (B-nZVI) by reduction of Fe3þ ions with NaBH4, and characterization of the produced material with SEM, XRD and BET-N2 technology; (2) evaluation of the factors impacting on Cr(VI) removal using an orthogonal method; nZVI and bentonite were used for Cr(VI) removal individually as a control, and the kinetics of Cr(VI) reduction by B-nZVI were also evaluated; and (3) remediation of electroplating wastewater including some heavy metal ions using B-nZVI and evaluation of reuse.
2.
Materials and methods
2.1.
Materials and chemicals
Bentonite was provided by Fenghong Co. Ltd, Anji, Zhejiang, China, primarily as Na-Mt montmorillonite (>90%), the chemical composition was 62.5% SiO2, 18.5% Al2O3, 1.75% Fe2O3, 4.25% MgO, 0.95% CaO, and 2.75% Na2O. The cation exchange capacity (CEC) was 75.4 meq/100g. After drying overnight at 80 C, the raw bentonite was ground and sieved through a 200 mesh screen prior to use in experiments. All the reagents were analytical grade (Shanghai Nanxiang Reagent Co., Ltd., China) and distilled water was used in all preparations. A stock solution containing potassium dichromate (K2Cr2O7) was prepared by dissolving K2Cr2O7 with deionized water and a series of solutions used during the experiment were prepared by diluting the stock to the desired concentrations.
2.2.
Synthesis of nZVI and supported nZVI
The nZVI and B-nZVI were prepared using conventional liquid-phase methods via the reduction of ferric iron by borohydride without or with bentonite as a support material (Celebi et al., 2007). Bentonite (2.00 g) was initially placed into a three-necked open flask, and a ferric solution produced by dissolving ferric chloride hexa-hydrate (9.66 g) in an ethanolwater solution (50 mL, 4:1 v/v) was added and stirred for 10 min. Subsequently, a freshly prepared NaBH4 solution
887
(3.54 g of NaBH4 in 100 mL) was added drop-wise into the mixture with constant stirring for 20 min after addition. The whole process described above was performed under a N2 atmosphere with vigorous stirring to avoid the oxidization of B-nZVI. The formed suspension was filtered and the black nanoscale precipitate was washed three times with pure ethanol and dried overnight at 75 C under vacuum (Celebi ¨ zu¨m et al., 2009). The theoretical mass fraction et al., 2007; U of bentonite in synthesized B-nZVI was 50%, and nZVI was prepared under identical conditions but with bentonite omitted. The nZVI and supported nZVI samples were stored in brown, sealed bottles under dry conditions and were not acidified prior to use.
2.3.
Characterizations and measurements
Scanning electron microscopy (SEM) was performed using a Philips-FEI XL30 ESEM-TMP (Philips Electronics Co., Eindhoven, The Netherlands). Images of various materials were obtained at an operating voltage of 30 kV. The SSA of nZVI, B-nZVI, and bentonite was measured via the BET adsorption ¨ zu¨m et al., 2009) using Micromeritics’ ASAP 2020 method (U Accelerated Surface Area and Porosimetry Analyzer (Micromeritics Instrument Corp.,USA). The specific surface areas of nZVI, B-nZVI and bentonite were 54.04, 39.94 and 6.03 m2/g, respectively. X-ray diffraction (XRD) patterns of B-nZVI before and after contacting Cr(VI) were performed using a PhilipsX’Pert Pro MPD (Netherlands) with a high-power Cu- Ka radioactive source (l ¼ 0.154 nm) at 40 kV/40 mA. The concentration of total Cr in solution and the concentrations of different heavy metal ions in the wastewater were determined using a flame atomic absorbance spectrometer (VARIAN AA 240FS, USA), and the Cr(VI) concentration was determined using the 1,5-diphenylcarbazide method (Geng et al., 2009) on a 722N visible spectrophotometer (Shanghai Precision & Scientific Instrument Co., Ltd, China).
2.4.
Batch experiments
The orthogonal method was used to test the effects of various factors on the reaction, and to optimize the conditions for Cr(VI) removal using B-nZVI. The experimental design was developed with the aid of the Orthogonal Design Assistant, where the initial concentration of Cr(VI), B-nZVI loading, temperature and pH were chosen as variables. Cr(VI) solutions (25 mL) with a known mass of B-nZVI were sealed in 50 mL centrifuge tubes and mixed for 4h before being centrifuged prior to analysis of the aqueous phase for residual Cr(VI). In order to investigate the role that bentonite and Fe0 played in the B-nZVI system, nZVI, B-nZVI and bentonite were all used in batch experiments examining Cr(VI) removal from aqueous solutions at an initial concentration of 50 mg/L at 35 C and 250 r/min. As the mass ratio of Fe0:bentonite was 1:1 in the B-nZVI system, the dosages of nZVI and bentonite were both set at 1.5 g/L, which was half the B-nZVI dosage. The mixtures were filtered through 0.45 mm mixed cellulose ester (MCE) membranes prior to determining the residual concentrations of Cr(VI) after contacting for 3 h. In order to investigate the effects of the different factors mentioned in
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the orthogonal experiments further, the mixtures of Cr(VI) solution and B-nZVI were mixed in 50 mL centrifugal tubes in a rotary shaker for determined periods of time, the conditions of which were initially set at 25 mL of 50 mg/L Cr(VI) solution, 3 g/L of B-nZVI, 35 C and 250 r/min. At selected timed intervals, the suspension was filtered through 0.45 mm MCE membranes, and the concentration of Cr(VI) in the filtrate was determined. To explore the feasibility of removing heavy metal ions from wastewater, B-nZVI was used to remediate electroplating wastewater collected from an electroplating factory’s sewage outfall (Fuzhou, China). The wastewater was centrifuged at 3000 r/min for 10 min to remove any insoluble impurities, prior to determining the initial pH and concentrations of total Cu, Cr, Pb and Zn. A batch of 50 mL bottles containing wastewater (10 mL) and B-nZVI (0.10 g) were mixed on a rotary shaker at 35 C and 250 r/min for 4 h. Then the mixtures were centrifuged at 3000 r/min for 10 min and the upper aliquot collected to determine pH and the concentration of each heavy metal ion. The potential to reuse B-nZVI for removing Cr(VI) from aqueous solution was also evaluated. B-nZVI (0.1 g) was added to 50 mg/L Cr(VI) solution (25 mL) and the mixture was shaken on a rotary shaker (35 C and 250 r/min). After 3 h, centrifugations at 3000 r/min were performed for 10 min to obtain solid-liquid separation. The supernatant was decanted carefully and used to determine the concentration of Cr(VI) remaining in solution while the used B-nZVI was mixed with different concentrations of EDTA. The residual B-nZVI-EDTA solution was washed with distilled water three times and shaken for another 3 h under identical conditions. The B-nZVI treated with 50 mg/L and 10 mg/L of EDTA was used to remove Cr(VI) for 3 times in succession to test the efficacy of reuse. In order to ascertain the accuracy, reliability and reproducibility of the data, orthogonal experiments were conducted in quadruplicate (n ¼ 4) and other batch experiments were carried out in triplicate (n ¼ 3) to minimize any experimental errors. The average values of the parallel measurements were used in all analysis and together with the standard deviations of these means were listed in Tables 1 and 3.
3.
Results and discussion
3.1.
Characterization
Table 1 e Orthogonal experimental design and the results obtained from the full 24 factorial experiment matrix. Removal T Cr(VI)ina pHina B-nZVIina Cr(VI)res (mg/L) (mg/L) efficiency (%) ( C) (mg/L) 1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16
25 25 25 25 30 30 30 30 35 35 35 35 40 40 40 40
20 50 70 100 20 50 70 100 20 50 70 100 20 50 70 100
2.0 5.0 8.0 10 5.0 2.0 10.0 8.0 8.0 10.0 2.0 5.0 10.0 8.0 5.0 2.0
2 3 4 5 4 5 2 3 5 4 3 2 3 2 5 4
0.11a 29 1 45.0 0.6 73 2 0a 0.06a 55.3 0.9 75 2 0.04a 31.1 0.7 0.04a 71 2 0.10a 6.75a 32.0 0.9 0.24a
99.5 42.0 35.7 26.6 100 99.9 21.0 25.3 99.8 55.6 99.9 29.4 99.5 86.5 54.3 99.8
ina - initial; res - residual; a means the standard deviations are too low to be listed.
presence of bentonite (Fig. 1). The synthesized nZVI without bentonite as a support material showed that nZVI particles were roughly globular and aggregated into a chain-like conformation (Fig. 1a). The diameters of the nanoscale zerovalent iron particles were in the range of 20e90 nm when bentonite was introduced as a support material. Compared with Fig. 1a, the aggregation of nZVI particles seemed to decrease and their dispersity increase in Fig. 1b, where the mass fraction of bentonite was 50%. A similar conclusion has been drawn using kaolin as a support material to synthesize kaolin supported nZVI, which was used to remove Cu(II) and Co ¨ zu¨m et al., 2009). As indicated in (II) from an aqueous solution (U Fig. 1c, the sizes of iron nanoparticles increase prominently after reacting with Cr(VI). This phenomenon could be attributed to the co-precipitation of Cr(III) and Fe(III) on the surface of the nanoparticles (Ponder et al., 2000; Manning et al., 2007), which occurs due to a redox reaction between Cr(VI) and Fe0 (Ponder et al., 2000; Manning et al., 2007). The XRD patterns of synthesized materials were compared with the XRD patterns obtained from standard materials, to identify the apparent peaks attributable to different iron and chromium compounds. The XRD patterns of B-nZVI before reaction (Fig. 2a) with Cr(VI) showed an apparent peak of Fe0 (2q z 44.90), which weakened significantly after the reaction
The SEM images of nZVI and B-nZVI showed the morphology and nanoparticle distribution of nZVI in the absence or
Table 2 e Range analysis and variance analysis of the orthogonal test. Factors
Temperature ( C) Cr(VI)ina (mg/L) pHina B-nZVIina (mg/L)
Range Analysis
Variance Analysis
k1
k2
k3
k4
Ranges
SSE
DOF
F-value
F critical values
50.9 99.7 99.7 59.1
61.6 71.0 56.4 66.7
71.2 52.7 61.8 72.8
85.0 45.3 50.7 70.1
34.1 54.4 49.1 13.7
2518 7039 5912 422
3 3 3 3
0.63 1.77 1.49 0.11
3.49 3.49 3.49 3.49
ina - initial; SSE - the Square Sum of Errors; DOF - Degree of Freedom.
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optimal conditions for chromium removal were 40 C, 20 mg/L of initial Cr(VI) concentration, 4 g/L B-nZVI loading and pH 2.0.
Table 3 e Remediation of actual electroplating wastewater by B-nZVI. Wastewater
Total Cr
Pb2þ
Cu2þ
c0 (mg/L) 73 2 13.1 0.6 33 1 a 2.4 0.1 c’ (mg/L) after reaction a Removal amount 7.3 1.3 3.0 (mg/g B-nZVI) Removal percentage (%) 100 100 92.7
Zn2þ
pH
284 4 1.9 115 3 4.5 16.8 e 59.4
e
a means the concentration of the heavy metal was under the limit of detection.
¨ zu¨m et al., 2009). The XRD patterns of B-nZVI after (Fig. 2b) (U reaction (Fig. 2b) indicated the presence of g-Fe2O3 (2q ¼ 35.68), Fe3O4 (2q ¼ 35.45) and Cr2FeO4 (2q ¼ 35.50), which were not detected in the sample before reaction (Chen et al., 2008). The appearance of Fe(II), Fe(III) and Cr(III) in B-nZVI after reaction demonstrated the occurrence of redox reactions between Fe0 and Cr(VI) where nZVI particles were acting as reductants, which was consistent with previous literature (Ponder et al., 2000).
3.2.
889
Orthogonal test
The designed complex conditions and the final removal efficiencies of Cr(VI) from aqueous solution by B-nZVI were listed in Table 1. Results were processed using the Orthogonal Design Assistant software (El Hajjouji et al., 2008) in Range Analysis and Variance Analysis (Table 2). The higher the range and the F-value, the more significant the factor was and the greater the influence of the factor on Cr(VI) removal Comparing the ranges and F-values in Table 2, factors influencing Cr(VI) removal were (in order of decreasing influencing): initial Cr(VI) concentration > pH > temperature > BnZVI loading. The K values from K1 to K4 represented each level of each factor, from the lowest to the highest. The higher the K value, the higher the removal efficiency, and the better the level of the factor. Take temperature for example, the highest K was K4, which represented the level 40 C, and this made 40 C the optimum temperature. The change in K values indicated that chromium removal increased with temperature and decreased as initial Cr(VI) concentration and pH rose. The
3.3.
Conditions affecting Cr(VI) removal
After contacting for 3 h under identical conditions, the removal efficiencies of Cr(VI) were 5.5, 60.0 and 100.0% respectively when bentonite, nZVI and B-nZVI were added individually. Bentonite generally has poor adsorption of Cr(VI) due to its negatively charged surface and the predominant existence of Cr(VI) as anions (Bhattacharyya and Gupta, 2008). The activity of nanoscale zero-valent iron particles was enhanced significantly when bentonite was introduced as a support material, which confirmed the role bentonite played as a dispersant and stabilizer in B-nZVI (Ponder et al., 2000). Kinetics studies of Cr(VI) reduction using B-nZVI suggested that the reactivity of nZVI particles supported on bentonite were enhanced significantly due to an increase in SSA and a decrease in aggregation. Reduction kinetics of Cr(VI) by B-nZVI were described by a pseudo first-order reaction (Ponder et al., 2000): dc v ¼ ¼ kSA as rm c dt
(1)
Where c was the concentration (mg/L) of contaminant, kSA was the specific reaction rate constant associated with the SSA of the materials (L/h m2), as was the specific surface area (m2/g), and rm was the mass concentration (g/L). For kSA, as, and rm are constant for a specific reaction, the product of the three can be expressed with one parameter kobs, which is called the observed rate constant of a pseudo first-order reaction (h1). Therefore Eq. (1) can be integrated into: ln
c ¼ kobs t c0
(2)
The kobs values under different conditions are equal to the slope of the line achieved by plotting lnðc=co Þ versus time under various conditions. In this study, the plots of lnðc=co Þ versus time produced linear plots with correlation coefficients (R2) higher than 0.9 (Fig. 3). This indicated that the rate of Cr(VI) reduction by B-nZVI fitted well the pseudo first-order model under various conditions. Additionally, the reduction of Cr(VI) by B-nZVI represented a solid-liquid inter-phase reaction, which agreed with the pseudo first-order kinetics model.
Fig. 1 e SEM images of laboratory synthesized iron particles with and without a support material. a. NZVI; b. B-nZVI before reaction with Cr(VI) solution; c. B-nZVI after reaction with Cr(VI) solution. The scale bar in the figure is 500 nm.
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w a t e r r e s e a r c h 4 5 ( 2 0 1 1 ) 8 8 6 e8 9 2
3.3.1.
Fig. 2 e X-ray diffractogram of B-nZVI. a. before the reaction with Cr(VI) solution; b. after the reaction with Cr (VI) solution.
Effect of initial Cr(VI) concentration
The effect of initial Cr(VI) concentration on removal efficiency was investigated in the range 20e100 mg/L. The plot fitted the pseudo first-order model well (Fig. 3a), where the observed rate constant decreased significantly as the initial Cr(VI) concentration increased, which agreed with the orthogonal test results. The equilibrium time became longer and the final removal efficiency of Cr(VI) decreased as the initial Cr(VI) concentration increased, so that while the percentage of Cr(VI) removed within 20 min at a Cr(VI) concentration of 20 mg/L was nearly 100%, it was only 30.4% within 60 min at a Cr(VI) concentration of 100 mg/L. Generally, the slower rate and lower efficiency of Cr(VI) removal from aqueous solution were found at higher concentrations of Cr(VI). Based on the SEM analysis and previous research, Cr(VI) reduction by nZVI could be defined as a surface-mediated process (Ponder et al., 2000; Rivero-Huguet and Marshall, 2009). The more the Cr(VI) ion approached the surface of nZVI dispersed on the bentonite, the faster Fe0 was oxidized into Fe(III) and the faster the coprecipitation of Cr(III) and Fe(III) oxides/hydroxides occurred. This reduced the reactivity of nZVI and subsequently resulted
Fig. 3 e Effects of various factors on Cr(VI) removal by fitting to the pseudo first-order model. a. initial Cr(VI) concentration; b. B-nZVI loading; c. pH value; d. temperature.
w a t e r r e s e a r c h 4 5 ( 2 0 1 1 ) 8 8 6 e8 9 2
in a decrease in kobs (Geng et al., 2009; Yuan et al., 2009). On the other hand, the highest removal amount was obtained at an initial Cr(VI) concentration of 70 mg/L, which could be ascribed to the limited capacity of B-nZVI for Cr(VI) removal determined by SSA.
3.3.2.
Effect of B-nZVI loading
The initial loadings of B-nZVI in Cr(VI) solution were 1, 2, 3 and 4 g/L. The Cr(VI) removal percentage increased as the B-nZVI concentration increased (Fig. 3b). The removal percentage of Cr(VI) was 54.6% using B-nZVI at 1 g/L for 120 min, but was nearly 100% when the B-nZVI loading was over 3 g/L. Meanwhile, kobs increased as the B-nZVI loading increased. These phenomena can be attributed to the increase in the available active sites resulting from the elevation in B-nZVI loading, where the reduction of Cr(VI) occurred (Geng et al., 2009; Yuan et al., 2009). However, the concentration of Cr(VI) decreased dramatically in the initial 10 min, then slightly declined in the later reduction. A few researchers (Ponder et al., 2000; Manning et al., 2007) reported a sorption phase during the reaction which could also be supported by our SEM images, suggesting that the overall mechanism was more complicated than a simple chemical reaction.
3.3.3.
Effect of the pH value
The two dominant forms of Cr(VI) in aqueous solution were 2 HCrO 4 , between pH 1.0 to 6.0 and CrO4 above pH 6.0 (Mohan and Pittman, 2006). The dependence of the reaction rate constant on pH was investigated by adjusting the solution pH to 2.0, 4.0, 6.0 and 8.0 with either 0.1 M HCl or NaOH (Yuan et al., 2009). Except for pH 2.0, the reduction of Cr(VI) can be described using the pseudo first-order model well (Fig. 3c). A remarkable increase in the removal rate occurred at pH 2.0, where equilibrium was achieved within 1 min and the residual Cr(VI) was below the detection limit, which made kinetic fitting infeasible. The Cr(VI) removal percentage decreased significantly with increases in the initial pH, so that only 27.2% Cr(VI) was reduced at pH 8.0 in 20 min while nearly 100% Cr(VI) was removed in 1 min at pH 2.0. In addition, kobs was respectively 0.0275, 0.0163, and 0.0083/min, when the initial pH value was 4.0, 6.0 and 8.0, indicating that the reduction rate increased as pH decreased, which had also been reported in other studies (Geng et al., 2009; Yuan et al., 2009). These results demonstrated that a lower pH favored Cr(VI) reduction, since at lower pH corrosion of nZVI was accelerated and the precipitation of Cr(III) and Fe(III) hydroxides on the surface of iron was consequently not as favorable, which led to an increase in the reaction rate (Lee et al., 2003). Furthermore, the increase in Hþ concentration left the surface of bentonite less negatively charged, which reduced the electrostatic repulsion between bentonite and Cr(VI) anions. This consequently promoted the electron transfer between zero-valent iron and Cr(VI) (Yuan et al., 2009).
3.3.4.
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ambient temperature. Fig. 3d highlights the relationship between lnðc=co Þ and time, where the linearity suggested that the reduction of Cr(VI) at different temperatures in the presence of B-nZVI fitted pseudo first-order dynamics (Ponder et al., 2000; Manning et al., 2007). The kobs was 0.020, 0.023, 0.026 and 0.030/min at four temperatures (25, 30, 35 and 40 C), showing that an increase in the reaction temperature resulted in an improved reaction rate. The apparent activation energy (Ea) of Cr(VI) reduction by B-nZVI was 24.9 kJ/mol, demonstrating that it is a chemically controlled adsorption process having an Ea value higher than 21 kJ/mol (Geng et al., 2009).
3.4. B-nZVI used to remove Cr(VI) from electroplating wastewater and B-nZVI reuse The data obtained from batch experiments where B-nZVI was used to remove Cr(VI) and other metals from electroplating wastewater are presented in Table 3, which indicated that BnZVI had the capacity to remove various heavy metals and was a potential promising candidate for applications to in situ environmental remediation. After reacting 10 mL of the wastewater with 0.1 g of B-nZVI for 4 h, the residual concentration of each metal ion showed that 100% total Cr, 100% Pb (II), 92.7% Cu(II), and 59.4% Zn(II) were removed, following treatment with B-nZVI. Total Cr, Pb(II) and Cu(II) received higher removal percentages due to their higher standard reduction potentials compared to Fe(II) (fqFeðIIÞ=Fe0 ¼ 0:44V). In contrast, a lower removal percentage of Zn(II) was obtained because the standard reduction potential of Zn(II) (fqZnðIIÞ=Zn0 ¼ 0:762V) was more negative than Fe(II) (Ladd, 2004). The amount of Cr(VI) removed when using B-nZVI treated with different concentrations of EDTA after being used four times was calculated and it was shown that B-nZVI’s ability to remove Cr(VI) was dramatically reduced after being used only once with an initial Cr(VI) concentration of 50 mg/L (Fig. 4). The rapid deterioration of B-nZVI was ascribed to the inability of the redox reaction between Cr(VI) and Fe0 to proceed
Effect of temperature
To assess the effect of different temperatures, batch experiments were conducted at 25, 30, 35 and 40 C. The results showed that 82.4% Cr(VI) was removed at 40 C while only 73.4% Cr(VI) was reduced at 25 C in 60 min. Thus zero-valent iron could have a positive effect on Cr(VI) reduction even at
Fig. 4 e The variation of Cr(VI) removal amount by B-nZVI after reusing four different times. The solutions of EDTA used for treatment of B-nZVI were 50 mg/L and 10 mg/L respectively as marked in the figure.
892
w a t e r r e s e a r c h 4 5 ( 2 0 1 1 ) 8 8 6 e8 9 2
further since this was a chemical controlled and irreversible process. As confirmed by XRD analysis (Fig. 2), reaction products were deposited on the surface of B-nZVI in the form of oxide-hydroxide co-precipitation of Fe(II), Fe(III) and Cr(III), which consequently decreased the activity of Fe0 (Chen et al., 2008). This phenomenon confirmed that the active ingredient of B-nZVI was Fe0 which acted as a reductant, while bentonite only played a role as a dispersant and stabilizer.
4.
Conclusions
In this study, nZVI particles became more effective when bentonite was introduced as a support material due to reduction of aggregation and increased SSA. Batch experiments indicated that the removal rate increased as the temperature and B-nZVI loading (4 g/L) increased, and fell as the initial Cr(VI) concentration and pH increased, which agreed with the result obtained from the orthogonal test. Under the various operational conditions considered, reduction of Cr(VI) using B-nZVI was in accordance with a pseudo first-order model. B-nZVI was effective in removing Cr(VI) and other heavy metals, including Pb(II), Cu(II) and Zn(II) from electroplating wastewater. Since bentonite is a stable and lowcost clay mineral, B-nZVI could be an efficient and promising remediation material to remove Cr(VI) and other metals from wastewater. However, further research must be carried out to slow and control the degree of nZVI oxidation in the atmosphere and as a consequence a more effective regeneration method may emerge from such studies.
Acknowledgements This work is supported by the Fujian “Minjiang Fellowship” from Fujian Normal University (gs1). The authors also gratefully acknowledge the significant contributions of Dr Gary Owens in editing and improving the many revisions of this manuscript, his suggestions and corrections have undoubtedly significantly improved the quality of the final manuscript.
references
Bhattacharyya, K.G., Gupta, S.S., 2008. Adsorption of a few heavy metals on natural and modified kaolinite and montmorillonite: a review. Adv. Colloid Interface Sci. 140, 114e131. ¨ zu¨m, C¸., Shahwan, T., Erten, H.N., 2007. A radiotracer Celebi, O., U study of the adsorption behavior of aqueous Ba2þ ions on
nanoparticles of zero-valent iron. J. Hazard. Mater. 148, 761e767. Chen, S.S., Hsu, B.C., Hung, L.W., 2008. Chromate reduction by waste iron from electroplating wastewater using plug flow reactor. J. Hazard. Mater. 152, 1092e1097. Cumbal, L., Greenleaf, J., Leun, D., SenGupta, A.K., 2003. Polymer supported inorganic nanoparticles: characterization and environmental applications. React. Funct. Polym. 54, 167e180. El Hajjouji, H., Ait Baddi, G., Yaacoubi, A., Hamdi, H., Winterton, P., Revel, J.C., Afidi, M., 2008. Optimisation of biodegradation conditions for the treatment of olive mill wastewater. Bioresour. Technol. 99, 5505e5510. Geng, B., Jin, Z., Li, T., Qi, X., 2009. Kinetics of hexavalent chromium removal from water by chitosan-Fe0 nanoparticles. Chemosphere 75, 825e830. Ju-Nam, Y., Lead, J., 2008. Manufactured nanoparticles: an overview of their chemistry, interactions and potential environmental implications. Sci. Total Environ. 400, 396e414. Katz, S., Salem, H., 1994. The Biological and Environmental Chemistry of Chromium. VCH Publishers, New York. Ladd, M.F.C., 2004. Introduction to Physical Chemistry, third ed. Cambridge University Press, Cambridge. Lee, T., Lim, H., Lee, Y., Park, J.W., 2003. Use of waste iron metal for removal of Cr(VI) from water. Chemosphere 53, 479e485. Li, X., Elliott, D.W., Zhang, W.X., 2006. Zero-valent iron nanoparticles for abatement of environmental pollutants: materials and engineering aspects. Crit. Rev. Solid State Mater. Sci. 31, 111e122. Li, Z., Kirk Jones, H., Zhang, P., Bowman, R.S., 2007. Chromate transport through columns packed with surfactant-modified zeolite/zero valent iron pellets. Chemosphere 68, 1861e1866. Liu, Y., Phenrat, T., Lowry, G.V., 2007. Effect of TCE concentration and dissolved groundwater solutes on NZVI-promoted TCE dechlorination and H2 evolution. Environ. Sci. Technol. 41, 7881e7887. Manning, B.A., Kiser, J.R., Kwon, H., Kanels, S.R., 2007. Spectroscopic investigation of Cr (III)-and Cr (VI)-treated nanoscale zerovalent iron. Environ. Sci. Technol. 41, 86e592. Mohan, D., Pittman, C.U., 2006. Activated carbons and low cost adsorbents for remediation of tri-and hexavalent chromium from water. J. Hazard. Mater. 137, 62e811. Ngomsik, A., Bee, A., Draye, M., Cote, G., Cabuil, V., 2005. Magnetic nano-and microparticles for metal removal and environmental applications: a review. Comptes RendusChimie 8, 963e970. Ponder, S.M., Darab, J.G., Mallouk, T.E., 2000. Remediation of Cr (VI) and Pb (II) aqueous solutions using supported, nanoscale zero-valent iron. Environ. Sci. Technol. 4, 2564e2569. Rivero-Huguet, M., Marshall, W.D., 2009. Reduction of hexavalent chromium mediated by micron-and nano-scale zero-valent metallic particles. J. Environ. Monit. 11, 1072e1079. ¨ zu¨m, C¸., Shahwan, T., Ero lu, A.E., Hallam, K.R., Scott, T.B., U Lieberwirth, I., 2009. Synthesis and characterization of kaolinite-supported zero-valent iron nanoparticles and their application for the removal of aqueous Cu2þ and Co2þ ions. Appl. Clay Sci. 43, 172e181. Yuan, P., Fan, M., Yang, D., He, H., Liu, D., Yuan, A., Zhu, J., Chen, T., 2009. Montmorillonite supported magnetite nanoparticles for the removal of hexavalent chromium [Cr (VI)] from aqueous solutions. J. Hazard. Mater. 166, 821e882.