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Effect of clay mineralogy on the feasibility of electrokinetic soil decontamination technology
Applied Clay Science 20 (2002) 283 – 293 www.elsevier.com/locate/clay
Effect of clay mineralogy on the feasibility of electrokinetic soil decontamination technology Darmawan a, S.-I. Wada b,* a
Department of Bioresources and Environmental Sciences, Graduate School of Kyushu University, Fukuoka 812-8581, Japan b Laboratory of Soils, Faculty of Agriculture, Kyushu University, Fukuoka 812-8581, Japan Received 18 January 2001; received in revised form 21 June 2001; accepted 3 July 2001
Abstract To evaluate the effect of clay mineralogy on the feasibility of electrokinetic soil remediation technology, we contaminated six soils with Cu(II), Zn(II) and Pb(II) and performed electroremediation for 570 h. Cation exchange resin saturated with H + was placed between soil and cathode to prevent soil alkalinization and trap the migrated heavy metal cations. After the treatment, the heavy metal cations were sequentially extracted with water, 1 M MgCl2 and hot 6 M HCl. In soils dominated by crystalline clay minerals, Cu(II) and Zn(II) significantly migrated from anode end and accumulated at the cathode end forming sparingly soluble hydroxides. Removal rates of Cu(II) and Zn(II) were highest in a soil dominated with kaolinite and crystalline hematite. In humic – allophanic and allophanic soils, the high pH-buffering capacity of allophane kept the soil pH above 5, even at the anode end, and Cu(II) and Zn(II) did not migrate significantly. In all soils, the migration of Pb(II) was infinitesimal due to the formation of insoluble PbSO4 and very strong surface complexation at the mineral surfaces. These results show that the reactivity of component clay minerals to H + and heavy metal cations has a crucial effect on the efficiency of the electrokinetic remediation technology and it is not effective for remediation of allophanic soils. The results also indicate that allophanic soils may be useful as a barrier material in landfill sites. D 2002 Elsevier Science B.V. All rights reserved. Keywords: Clay mineralogy; Electrokinetic remediation; Heavy metal; Soil pollution
1. Introduction The direct disposal, dumping and tailing of wastes containing toxic heavy metals often result in accumulation of the heavy metals in soils because they are strongly retained by some soil components such as clay minerals and humic substances. The heavy
*
Corresponding author. Tel.: +81-92-642-2844; fax: +81-92642-2864. E-mail address: [email protected] (S.-I. Wada).
metals retained in soils are, however, gradually released into pore water, resulting in pollution of surface and ground waters. In addition, fine-grained soil particles retaining heavy metals are transported as airborne dust and contaminate agricultural land and crops. Thus, the heavy metal-contamination of soils occurring in industrial areas is a great concern due to its direct and indirect harmful effects on human health. Among many technologies developed to decontaminate heavy metal-polluted soils, the electrokinetic method is regarded as an effective technique particularly for soils having low hydraulic conductivity
0169-1317/02/$ - see front matter D 2002 Elsevier Science B.V. All rights reserved. PII: S 0 1 6 9 - 1 3 1 7 ( 0 1 ) 0 0 0 8 0 - 1
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(Acar and Alshawabkeh, 1993). In electrokinetic remediation, a direct current is passed through a soil to induce the electromigration and electroosmosis towards the electrode where the contaminants are collected. According to an analysis by Acar and Alshawabkeh (1993), electromigration is more important at least for removal of ionic contaminants. For successful decontamination of heavy metal ions, therefore, it is important to convert the precipitated and adsorbed ions into dissolved forms. Heavy metals incorporated into soils take different chemical forms including (1) dissolved ionic form, (2) electrostatically adsorbed form, and (3) surface complexed form (Darmawan and Wada, 1999). Layer silicate minerals having excess surface negative charge arising from isomorphous substitution adsorb heavy metal cations electrostatically. Due to the nonspecific nature of the coulombic force, the adsorbed heavy metal cations are easily exchanged by other cations. The selectivity coefficient, for example, for Cu –Na exchange on a montmorillonite is near unity (Sposito et al., 1981). On the other hand, oxides and hydroxides of iron and aluminum as well as humic substances bind heavy metal cations very strongly, forming surface complexes, in which the bonding between heavy metal cation and surface functional groups bears some covalency. The surface complexed heavy metal cations are, therefore, not exchanged by common ions in the pore water of soils. Thus, it is expected that the feasibility of the electrokinetic remediation method depends strongly on min-
eralogical composition as well as on soil organic matter content. Studies by Puppala et al. (1997), Reddy et al. (1997), and Grundl and Reese (1997) suggested that the electrokinetic method is not necessarily effective for soils with high adsorption capacity and also for those containing calcium carbonate. To date, however, many researchers have used only artificially polluted model soils made up of pure clays such as kaolinite (Acar and Alshawabkeh, 1996; Yeung et al., 1996; Dzenitis, 1997) and smectite (Reddy et al., 1997; Grundl and Reese, 1997) or their mixture with quartz (Kawachi and Kubo, 1999) in developing and evaluating the electrokinetic remediation technology. There seems to be no systematic experimental study in which the electrokinetic remediation method is applied to a series of natural soils that differ in mineralogical composition. In the present study, therefore, six soils that differ in clay mineralogy and organic matter content were contaminated with salts of copper, lead, and zinc, subjected to electrokinetic remediation treatment, and the behaviors of the heavy metal cations were analyzed through successive extraction.
2. Materials Six soil samples from different locations were used in the present study. They were air-dried and passed through a 2-mm sieve. Some chemical and mineralogical properties determined by standard procedures (Page et al., 1982; Klute, 1986) are listed in Table 1.
Table 1 Selected properties of soil samples Soil name
pH
Organic carbon (g kg 1)
Clay (g kg 1)
Allophane and/ or imogolite (g kg 1)
DCBa Fe2O3 (g kg 1)
Major cation exchanger
Nakajo Harumachi
6.8 5.8
2.98 15.29
382 246
nmb nm
23.8 32.4
Fukuchiyama
6.1
8.81
445
nm
24.9
Pakchong Goshi Choyo
4.9 5.9 6.2
6.18 62.67 22.79
771 385 371
nm 232 304
85.8 54.0 35.5
smectite vermicullite, micaceous, kaolinite micaceous, vermicullite, kaolinite kaolinite, hematite allophane, Al – humus imogolite, allophane
a b
Dithionite – citrate – bicarbonate. No measurement.
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As shown in Table 1, the Nakajo, Fukuchiyama, Harumachi, and Pakchong soils are predominated by crystalline clay minerals. Among these four soil samples, effective CEC of the Pakchong soil is the lowest irrespective of its extremely high clay content, reflecting the predominance of kaolinite in its clay fraction. The Goshi and Choyo soils are similar in that the clay consists exclusively of allophane and imogolite but the Goshi soil has much higher organic matter content. Table 1 also shows that the amounts of Fe and Mn oxides dissolved by dithionite –citrate –bicarbonate (DCB) treatment (Mehra and Jackson, 1960). All the soil samples have fairly high amounts of DCB soluble iron oxide with a maximum of 85.8 g kg 1 for the Pakchong soil, but the amounts of DCB soluble Mn oxides are much lower. Electron microscopy (photos not presented) showed that the iron oxide in the Pakchong soil exists as fine pseudo-hexagonal hematite particles, whereas no discrete iron oxide particles was detectable in other soils. The soils were loaded with 1000 mg kg 1 of Cu(II) and Zn(II) and 500 mg kg 1 of Pb(II) by mixing the soil samples with the calculated amounts of CuSO45H2O, ZnSO47H2O and Pb(NO3)2 and aged for 1 year at field moisture contents. These three metal cations were employed because they are expected to behave differently in soils (Darmawan and Wada, 1999), i.e., Cu(II) and Pb(II) are highly specifically adsorbed by oxides and humic substances but Pb(II) tends to form sparingly soluble sulfates, and Zn(II) is preferred rather by layer silicate minerals.
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The detailed procedure of the sample preparation is given in Darmawan and Wada (1999). Since Cu(II) and Zn(II) were added as soluble sulfate, sulfate ion from these salts would have reacted with Pb(II) ion from Pb(NO3)2 to form sparingly soluble PbSO4 during aging. Although this makes the system complicated, the present combination of salts was selected because PbSO4 is one of the major forms of Pb(II) in heavily polluted soils.
3. Methods The apparatus was made from an open plastic box with dimensions of 220 55 55 mm by partitioning it into four compartments with nylon cloth and attaching an inlet for anolyte, as shown in Fig. 1. The polluted soil samples were packed into the largest central compartment and a pair of graphite electrodes was placed in the two compartments at the two ends. The weight of the packed soil was 320 g for the Nakajo soil, 355 g for the Fukuchiyama soil, 350 g for the Harumachi soil, 335 g for the Pakchong soil, 317 g for the Goshi soil, and 201 g for the Choyo soil. The compartment between the soil and electrode was packed with H-saturated cation exchange resin (Dowex 50W 8) to neutralize OH ion generated at the cathode and to trap heavy metal cations (Wada and Ryu, 1999). The inlet of the apparatus, packed with the soil and resin, was connected to a constant head device containing 10 mM NaCl solution to saturate the
Fig. 1. Schematic illustration of the experimental cell.
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interstitial pores of the soil. After saturation, the level of the anolyte was set 5 mm below the soil surface (Wada et al., 1999) and 20-V potential was applied with a stabilized power supply. Throughout the treatment, the electric current was monitored with an ammeter inserted in the circuit (Fig. 1). After 570 h, the apparatus was disconnected from the power supply and the soil column was cut into six slices of equal length, and air-dried. The soil pH was measured at a soil/water ratio = 1:2.5. Portions (2 g) of the air-dried samples were placed in polycarbonate centrifuge tubes and extracted sequentially with water, 1 M MgCl2 and 6 M HCl. Water and MgCl2 extractions were carried out with 20 ml of water and 16 ml of an MgCl2 solution by shaking for 1 h at room temperature. HCl extraction was performed by digesting with 10 ml of 30% H2O2 to dryness followed by refluxing with 16 ml of 6 M HCl (Asami and Kato, 1977) at boiling point. The extracted Cu(II), Pb(II) and Zn(II) were determined by atomic absorption spectroscopy (AAS) following the procedures described by Amacher (1996) and Reed and Martens (1996). The resin was collected quantitatively and packed in a column. The adsorbed metal cations were extracted by leaching with 1 M NH4NO3 and determined by AAS. All the measurements were carried out in duplicate and the results were averaged.
4. Results and discussion The measured electric currents are plotted against elapsed time in Fig. 2. In general, the electric current decreased rapidly in the first 50-h period followed by gradual decrease in the subsequent 150 h for all soils. The mode of variation and the magnitude of the current were, however, significantly different among soils. For the Nakajo, Fukuchiyama and Harumachi soils, the initial electric current was about 13 –19 mA and it monotonically decreased to the final range of 1.6– 2.7 mA. The trend was quite similar for the Goshi and Choyo soils but both the initial and final values were much lower, i.e., 5.5– 6.2 mA and 0.2 – 0.5 mA, respectively. For the Pakchong soil, on the other hand, the electric current
Fig. 2. Time course of electric current.
increased up to 28 mA after 7 h and rapidly dropped down to < 1 mA after 200 h. The higher initial values of the electric current are obviously due to the higher electrolyte concentration in the pore water. Since indigenous soluble salts are negligible in all soils and the amount of added heavy metal salts were the same, the cause of the lower initial current for humic –allophanic Goshi and allophanic Choyo soils would be that the cations and anions of the added salts were adsorbed on the soil materials to a larger extent (Wada, 1984). The cause of the rapid increase and immediate decrease is not clear but the higher initial electric current in the Pakchong soil supports the contention that the initial contents of water soluble Cu(II), Pb(II) and Zn(II) were the highest in the Pakchong soil (Figs. 3 – 5). The pH profiles along the soil column after the treatment as well as the initial pH values are presented in Fig. 6. Since H + ion was generated by the electrolysis reaction of water continuously at the anode and it migrated toward the cathode, soil pH significantly dropped, particularly in the section neighboring the anode. In the Nakajo, Fukuchiyama, Harumachi and Pakchong soils predominated by crystalline clay minerals, the pH of the soil section next to the anode was below 3. Among these four soils, the Pakchong soil was conspicuous and the pH was far below 3 in the first through fifth sections. In the Nakajo, Fukuchiyama, and Harumachi soils, on the other hand, the magnitude of the pH drop decreased toward the cathode and the pH value of the soil section neighboring the cathode was quite near
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Fig. 3. Distribution of three forms of Cu(II) in soils after treatment.
the initial one. The magnitude of pH drops was much smaller in the humic – allophanic Goshi and allophanic Choyo soils and the pH was mostly kept above 5 except in the section neighboring the anode. The plots in Fig. 2 were interpolated by a cubic spline function and integrated over the treatment time to calculate the amount of the electric charge transported through the soil samples. The calculated amount of charge was 7.95 kC for the Nakajo, 5.72 kC for Fukuchiyama, 4.90 kC for Harumachi, 4.76 kC for Pakchong, 1.98 kC for Goshi, and 0.649 kC for Choyo soils. If it is assumed that the electrolytic generation of H + ion is proportional to the electric current, the H + ion supplies in the Goshi and Choyo soils were estimated to be much smaller than that in other soils. However, the pH profiles for the Goshi and Choyo soils are quite similar irrespective of the difference in the amount of the transported charge, and the differences in H + ion concentration estimated from the pH values are much smaller than those expected from the differences in the charge transport
(Fig. 2). These data suggest that the lower electric currents in the Goshi and Choyo soils are not a major cause for their higher pH values. Alternatively and probably the most important cause for that would be the higher acid-buffering capacity of the soils. Allophane and imogolite present in these soils contain large amounts of surface aluminol groups (Wada, 1984). On addition of an acid, HA, the aluminol groups uptake H + ion and the resulting positively charged sites retain the accompanying anion: Al-OH þ Hþ þ A ¼ Al-OHþ 2A :
ð1Þ
With this reaction, the H + ion generated at the anode would have largely been consumed, and in turn, the pH of the interstitial water dropped to a lesser extent. It seems contradictory that the Pakchong soil containing the largest amount of free iron oxide, which can uptake H + ions and anions, exhibited the lowest pH. However, iron oxide in the Pakchong soil was mostly crystalline hematite, whereas it was non-
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Fig. 4. Distribution of three forms of Pb(II) in soils after treatment.
crystalline in other soils. Higher crystallinity of free iron oxide and the predominance of kaolinite with low cation adsorbing capacity would have favored the acidification in the Pakchong soil. The contents of three forms of Cu(II), Pb(II), and Zn(II) in soil slices after electrokinetic treatment are presented in Figs. 3 – 5, together with their initial concentrations. Comparison of these figures indicates that the distribution pattern of Cu(II) and Zn(II) are similar for all soils while that of Pb(II) differed significantly. The distribution of Pb(II) was distinctly different in that the removal rate was significantly lower and the proportion of the MgCl2-extractable fraction was comparable to or even higher than the HCl-extractable fraction (Fig. 4). In the present extraction procedure, water-extractable Cu(II), Pb(II) and Zn(II) would come mostly from soluble salts of these heavy metals. MgCl2extractable fraction would basically represent those retained on layer silicate clay minerals as exchangeable cations. HCl-extractable fraction would corre-
spond to those bound as surface complexes on iron oxides, allophane and humic substances (Darmawan and Wada, 1999) as well as hydroxide precipitates. In addition, for Pb, the MgCl2-extractable fraction would include Pb(II) precipitated as PbSO4 (anglesite) in addition to exchangeable Pb(II), due to the increased solubility of PbSO4 in a concentrated Mg Cl2 solution. This occurs because the activity of sulfate ion is reduced by the formation of soluble MgSO4 complex. The ion speciation based on Pitzer’s model (Gueddari et al., 1983) showed that the ion activity product of (Pb2 + )(SO42 ) would never exceed 10 8 during MgCl2 extraction. Since the solubility product of PbSO4 is 1.62 10 7 (Lindsay, 1979), the precipitate of PbSO4 is expected to have dissolved during extraction and the MgCl2-extractable fraction can be regarded as a sum of exchangeable Pb and PbSO4. Before treatment, copper was found mostly in the strong acid-extractable fraction, except in the Pakchong soil. After the treatment, Cu(II) content re-
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Fig. 5. Distribution of three forms of Zn(II) in soils after treatment.
duced markedly in soil sections near the anode to less than 100 mg kg 1 and accumulated in sections near the cathode in the Nakajo, Fukuchiyama, and Harumachi soils. Comparison of Figs. 3 and 6 suggests that the amount of Cu(II) left in the soil is
Fig. 6. Profile of soil pH after treatment.
inversely correlated with pH. The stability of surface complexes of Cu(II) on oxide and humic substances decreases steeply below pH 4 (McKenzie, 1980; Kendorf and Schnitzer, 1980). The distribution pattern of Cu(II) shown in Fig. 3 suggests that Cu(II) was released from surface functional groups, migrated toward the cathode and hydrolyzed to precipitate when it entered into a region of higher pH, as repeatedly reported by many researchers (Yeung et al., 1996; Puppala et al., 1997; Viadero et al., 1998). A noticeable feature in Fig. 3 is that significant amount of MgCl2-extractable, i.e., exchangeable, Cu(II) appeared in the fourth to sixth sections from the anode in the Nakajo, Fukuchiyama, and Harumachi soils. Darmawan and Wada (1999) showed that heavy metal cations loaded to soils are rapidly adsorbed at permanent negative charge on layer silicate minerals, and then slowly transferred to oxides and humic substances to form surface complexes in 50 days. This and the results presented in Fig. 3 suggest that the role of cation exchange sites on
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layer silicate minerals becomes more important in a nonequilibrium migration process, particularly at low pH where the stability of surface complexes is low. Thus, longer treatment time and more electrical power are required for remediation of soils with high cation exchange capacity (Puppala et al., 1997). Fortunately, the cation exchange sites of layer silicate minerals do not show high preference for heavy metal cations (Sposito et al., 1981). The efficiency of electrokinetic remediation may be improved by adding salts of noncontaminant cations, e.g., CaCl2, that compete with heavy metal cations for the exchange site. In the Pakchong soil, the Cu(II) removal rate was the highest among the soils. Total content of the three forms of Cu(II) was below 100 mg kg 1 in the first and second sections and it was about 200 mg kg 1 in the fifth section, i.e., the removal rate was > 80%. The relatively high removal rate would have resulted from the relatively low pH of the original soil and low cation exchange capacity (Table 1) and higher crystallinity, and therefore, lower reactivity of iron oxide minerals. In contrast, Cu(II) was not removed at all from the Goshi soil (Fig. 3). It is apparent that a major cause of this is the high acid-buffering capacity of the soil (Fig. 6). In addition, the extremely high selectivity of humic substances for Cu(II) would have contributed the result. A significant reduction of Cu(II) content in the section neighboring the anode in the allophanic Choyo soil that showed the same pH value to the corresponding section of the humic –allophanic Goshi soil (Fig. 6) supports this view. Fig. 4 shows that the behavior of Zn(II) is basically similar to that of Cu(II) but Zn(II) is much more mobile under the applied electric field. The behavior of Pb(II) shown in Fig. 5 is distinctly different from that of Cu(II) and Zn(II). First, the proportion of MgCl2-extractable fraction is apparently larger in soils dominated by layer silicate minerals before treatment. This must be due to the presence of PbSO4 and its dissolution in 1 M MgCl2 solution as already discussed. Second, the removal rate was significantly lower than that for Cu(II) and Zn(II) and half or more of the remaining Pb(II) was in an MgCl2-extractable form in all soils, except for humic – allophanic and allophanic soils.
Since PbSO4 precipitated in soils does not move either via electrophoresis or via electroosmosis, lower removal rates were expected in soils in which a large proportion of Pb(II) was in a form of PbSO4. Actually the results obtained for the Nakajo, Fukuchiyama, Harumachi and Pakchong soils (Fig. 5) follow the expected trend. In these soils the HCl-extractable fraction, i.e., surface complexed Pb(II), reduced significantly in sections near the anode and seems to have accumulated in the following sections. In contrast, most of the Pb(II) was found in HClextractable fraction and the proportion of MgCl2extractable fractions were low in the Goshi and Choyo soils both before and after the treatment (Fig. 5). This indicates that the precipitation of PbSO4 was greatly suppressed in these soils. Possible mechanism of this is the simultaneous SO42 and Pb2 + adsorption. As already discussed, Pb2 + forms stable surface complexes with surface hydroxyl groups on minerals and carboxyl groups on humic substance. The reaction can be expressed, for example, 2ðSOHÞ þ Pb2þ ¼ ðSOÞ2 Pb þ 2Hþ ;
ð2Þ
where SOH stands for surface hydroxyl groups and carboxyl groups. In acidic condition, some surface hydroxyl groups, typically AlOH and FeOH groups, are protonated and adsorb anions. Among anions, oxo-anions are adsorbed more strongly. The reaction can be expressed as: 2ðAlOHÞ þ 2Hþ þ SO2 4 ¼ ðAlOH2 Þ2 SO4 :
ð3Þ
Since allophanic soils have large amount of the both types of functional groups, the overall reaction is: 2ðSOHÞ þ 2ðAlOHÞ þ Pb2þ þ SO2 4 ¼ ðSOÞ2 Pb þ ðAlOH2 Þ2 SO4 :
ð4Þ
With the simultaneous adsorption of Pb2 + and SO42 , the activity of these ions in solution would be maintained low enough to prevent formation of PbSO4(s). In the present study, H-saturated cation exchange resin was placed between cathode and soil to prevent alkalinization of soil and to trap the migrated heavy metal cations. Wada and Ryu (1999) have examined
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Table 2 Mass balance of Cu, Pb and Zn Heavy metal
Soil name
Initial (g)
Remained in soil
Adsorbed in resin
(g)
(%)
(g)
(%)
Total detected (%)
Cu
Nakajo Harumachi Fukuchiyama Pakchong Goshi Choyo
0.29 0.336 0.326 0.307 0.326 0.181
0.264 0.269 0.245 0.162 0.313 0.170
91.3 80.0 75.2 52.9 96.1 94.3
0.026 0.050 0.070 0.146 0.001 0.001
9.2 14.9 21.5 47.6 0.4 0.6
100.5 95.0 96.8 100.5 96.5 94.9
Pb
Nakajo Harumachi Fukuchiyama Pakchong Goshi Choyo
0.142 0.172 0.170 0.166 0.166 0.095
0.138 0.163 0.149 0.160 0.157 0.092
96.5 94.6 88.0 96.6 94.4 96.7
0.004 0.005 0.007 0.009 0.001 0.000
2.8 3.1 4.1 5.8 0.1 0.2
99.4 97.8 92.2 102.5 94.6 96.9
Zn
Nakajo Harumachi Fukuchiyama Pakchong Goshi Choyo
0.313 0.364 0.347 0.279 0.315 0.173
0.227 0.169 0.091 0.047 0.282 0.173
72.4 46.4 26.1 17.1 89.5 99.8
0.077 0.174 0.217 0.213 0.004 0.005
24.8 47.9 62.6 76.6 1.4 3.0
97.3 94.3 88.7 93.8 91.0 102.8
this in Cu(II) removal from an artificially contaminated soil and found that the alkalinization of the soil neighboring the anode was successfully suppressed and the whole soil was acidified. The pH profile in Fig. 6, however, shows that the alkalinization was prevented but the acid front did not reach the cathodic end of the soil. Therefore, the migrated heavy metal cations hydrolyzed and accumulated in the soil section nearest to the cathode. The reason for the difference between the present result and those by Wada and Ryu (1999) is not clear. One possible reason is that there was no drainage of catholyte in the present study. The mass balances of Cu(II), Zn(II) and Pb(II) are listed in Table 2. The percentages of the total detected heavy metals were mostly >90 except for Zn in the Fukuchiyama soil. Since there was no noticeable interference in the AAS determination, the missing portions of the heavy metals would have associated with the nylon cloth used for separating the resin and soils. Although there are some uncertainties, Table 2 clearly shows that the percentages of Cu(II) and Pb(II) trapped in the cation exchange resin were generally low for all the soil samples than those of Zn(II). Since the objective of the present study was to examine the effect of clay mineralogy on the efficiency of elec-
trokinetic remediation technology, any additional enhancement was not performed. With some enhancement techniques including addition of salts of noncontaminant cations, soil acidification, and drainage of catholyte, the removal rate would be improved, at least, for nonallophanic soils. The results for the Goshi and Choyo soils suggest that the electrokinetic technology would not work for humic – allophanic and allophanic soils, at least for Cu(II) removal. Nevertheless, this may not be a negative finding. The extremely low mobility of heavy metal cations in these soils even under electrokinetic treatment strongly indicates that heavy metal cations incorporated in these types of soils are expected not to be leached into the ground water under the conditions normally encountered in nature. This indicates the possible use of these types of soils for low-cost permeable barrier at waste landfill site.
5. Conclusions . Cu(II) and Zn(II) electrically migrated from the anode toward the cathode to a significant extent in
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soils dominated by crystalline clay minerals. But a lesser drop in pH of soil section near the cathode has led to the precipitation of Cu(II) and Zn(II) into hydroxides. Among these soils, the electrokinetic removal of Cu(II) and Zn(II) in the kaolinitic Pakchong was the highest due to low initial pH, low cation exchange capacity, and lower reactivity of iron oxide minerals of this soil. . The removal rates of Cu(II) and Zn(II) from soils dominated by crystalline clay minerals would be increased by some enhancement techniques, including addition of salts of competing noncontaminant cations, soil acidification, and catholyte drainage. . The high pH-buffering capacities of humic – allophanic and allophanic soils are the major factors that have hindered the electromigration of Cu(II) and Zn(II) in these soils. . Precipitation as sparingly soluble PbSO4 and stable surface complexation with hydroxyl groups of minerals and carboxyl groups of humic substances have resulted in extremely low migration of Pb(II) from all soils under study. . Strong retention of heavy metal cations by humic – allophanic and allophanic soils and their extremely low mobility even under electrical field indicate the possible use of these soils as an ideal permeable barrier at landfill site.
Acknowledgements We thank Dr. Mochizuki of Experimental Farm of Kyushu University for providing a soil sample. This study was supported in part by a Grant-in-Aid for Scientific Research (#11660066) from the Japanese Society for Promotion of Sciences.
References Acar, Y.B., Alshawabkeh, A.N., 1993. Principles of electrokinetic remediation. Environ. Sci. Technol. 27, 2638 – 2647. Acar, Y.B., Alshawabkeh, A.N., 1996. Electrokinetic remediation: I. Pilot-scale tests with lead-spiked kaolinite. ASCE J. Geotech. Eng. 122, 173 – 185. Amacher, M.C., 1996. Nickel, cadmium and lead. In: Sparks, D.L. (Ed.), Methods of soil analysis: Part 3. Chemical methods. Soil Science Society of America, Madison, WI, pp. 739 – 768.
Asami, T., Kato, K., 1977. Comparison of analytical methods for total cadmium, zinc, lead and copper in soils. Jpn. J. Soil Sci. Plant Nutr. 48, 335 – 336 (in Japanese). Darmawan, Wada, S.-I., 1999. Kinetics of speciation of Cu, Pb and Zn loaded to soils that differ in cation exchanger composition at low moisture content. Commun. Soil Sci. Plant Anal. 30, 2363 – 2375. Dzenitis, J.M., 1997. Soil chemistry effects and flow prediction in electroremediation of soil. Environ. Sci. Technol. 31, 1191 – 1197. Grundl, T., Reese, C., 1997. Laboratory study of electrokinetic effects in complex natural sediments. J. Hazard. Mater. 55, 187 – 201. Gueddari, M., Monnin, C., Perret, D., Fritz, B., Tardy, Y., 1983. Geochemistry of brines of the Chot El Jerid in southern Tunesia—application of Pitzer’s equations. Chem. Geol. 39, 165 – 178. Kawachi, T., Kubo, H., 1999. Model experimental study on the migration behavior of heavy metals in electrokinetic remediation process for contaminated soil. Soil Sci. Plant Nutr. 45, 259 – 268. Kendorf, H., Schnitzer, M., 1980. Sorption of metals on humic acid. Geochim. Cosmochim. Acta 44, 1701 – 1708. Klute, A., 1986. Methods of soil analysis: Pt. 2. Physical and mineralogical methods, 2nd edn. Soil Science Society of America, Madison, Wisconsin. Lindsay, W.L., 1979. Chemical equilibria in soils. Wiley, New York. McKenzie, R.M., 1980. The adsorption of lead and other heavy metals on oxides of manganese and iron. Aust. J. Soil Res. 18, 61 – 73. Mehra, O., Jackson, M.L., 1960. Iron oxide removal from soils and clays by a dithionite – citrate system with sodium bicarbonate buffer. Clays Clay Miner. 7, 313 – 327. Page, A.L., Miller, R.H., Keeney, D.R., 1982. Methods of soil analysis: Pt. 1. Chemical and microbiological properties, 2nd edn. Soil Science Society of America, Madison, Wisconsin. Puppala, S.K., Alshawabkeh, A.N., Acar, Y.B., Gale, R.J., Bricka, J., 1997. Enhanced electrokinetic remediation of high sorption capacity soil. J. Hazard. Mater. 55, 203 – 220. Reddy, K.R., Parupudi, U.S., Devulapalli, S.N., Xu, C.Y., 1997. Effects of soil composition on the removal of chromium by electrokinetics. J. Hazard. Mater. 55, 135 – 158. Reed, S.T., Martens, D.C., 1996. Copper and zinc. In: Sparks, D.L. (Ed.), Methods of soil analysis: Part 3. Chemical methods. Soil Science Society of America, Madison, WI, pp. 703 – 722. Sposito, G., Holtzclaw, K.M., Johnston, C.T., LeVesque-Madore, C.S., 1981. Thermodynamics of sodium – copper exchange on Wyoming bentonite at 298 K. Soil Sci. Soc. Am. J. 45, 1079 – 1083. Viadero Jr., R.C., Reed, B.E., Berg, M., Ramsey, J. 1998. A laboratory-scale study of applied voltage on the electrokinetic separation of lead from soils. Sep. Sci. Technol. 33, 1833 – 1859. Wada, S.-I., 1984. Mechanism of apparent salt absorption in Ando soils. Soil Sci. Plant Nutr. 30, 77 – 83.
Darmawan, S.-I. Wada / Applied Clay Science 20 (2002) 283–293 Wada, S.-I., Ryu, T., 1999. Electrokinetic remediation of Cu contaminated natural soil by using cation exchange resin. Proc. 34th Jpn. Conf. Geotech. Eng. 2, 2213 – 2214. Wada, S.-I., Ryu, T., Darmawan, Umegaki, Y., 1999. On the design of laboratory scale apparatus for electrokinetic soil decontami-
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nation under open-flow condition. J. Fac. Agric., Kyushu Univ. 43, 479 – 487. Yeung, T.A., Hsu, C., Menon, R.M., 1996. EDTA-enhanced electrokinetic extraction of lead. ASCE J. Geotech. Eng. 122, 666 – 673.
Effect of clay mineralogy on the feasibility of electrokinetic soil decontamination technology Darmawan a, S.-I. Wada b,* a
Department of Bioresources and Environmental Sciences, Graduate School of Kyushu University, Fukuoka 812-8581, Japan b Laboratory of Soils, Faculty of Agriculture, Kyushu University, Fukuoka 812-8581, Japan Received 18 January 2001; received in revised form 21 June 2001; accepted 3 July 2001
Abstract To evaluate the effect of clay mineralogy on the feasibility of electrokinetic soil remediation technology, we contaminated six soils with Cu(II), Zn(II) and Pb(II) and performed electroremediation for 570 h. Cation exchange resin saturated with H + was placed between soil and cathode to prevent soil alkalinization and trap the migrated heavy metal cations. After the treatment, the heavy metal cations were sequentially extracted with water, 1 M MgCl2 and hot 6 M HCl. In soils dominated by crystalline clay minerals, Cu(II) and Zn(II) significantly migrated from anode end and accumulated at the cathode end forming sparingly soluble hydroxides. Removal rates of Cu(II) and Zn(II) were highest in a soil dominated with kaolinite and crystalline hematite. In humic – allophanic and allophanic soils, the high pH-buffering capacity of allophane kept the soil pH above 5, even at the anode end, and Cu(II) and Zn(II) did not migrate significantly. In all soils, the migration of Pb(II) was infinitesimal due to the formation of insoluble PbSO4 and very strong surface complexation at the mineral surfaces. These results show that the reactivity of component clay minerals to H + and heavy metal cations has a crucial effect on the efficiency of the electrokinetic remediation technology and it is not effective for remediation of allophanic soils. The results also indicate that allophanic soils may be useful as a barrier material in landfill sites. D 2002 Elsevier Science B.V. All rights reserved. Keywords: Clay mineralogy; Electrokinetic remediation; Heavy metal; Soil pollution
1. Introduction The direct disposal, dumping and tailing of wastes containing toxic heavy metals often result in accumulation of the heavy metals in soils because they are strongly retained by some soil components such as clay minerals and humic substances. The heavy
*
Corresponding author. Tel.: +81-92-642-2844; fax: +81-92642-2864. E-mail address: [email protected] (S.-I. Wada).
metals retained in soils are, however, gradually released into pore water, resulting in pollution of surface and ground waters. In addition, fine-grained soil particles retaining heavy metals are transported as airborne dust and contaminate agricultural land and crops. Thus, the heavy metal-contamination of soils occurring in industrial areas is a great concern due to its direct and indirect harmful effects on human health. Among many technologies developed to decontaminate heavy metal-polluted soils, the electrokinetic method is regarded as an effective technique particularly for soils having low hydraulic conductivity
0169-1317/02/$ - see front matter D 2002 Elsevier Science B.V. All rights reserved. PII: S 0 1 6 9 - 1 3 1 7 ( 0 1 ) 0 0 0 8 0 - 1
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(Acar and Alshawabkeh, 1993). In electrokinetic remediation, a direct current is passed through a soil to induce the electromigration and electroosmosis towards the electrode where the contaminants are collected. According to an analysis by Acar and Alshawabkeh (1993), electromigration is more important at least for removal of ionic contaminants. For successful decontamination of heavy metal ions, therefore, it is important to convert the precipitated and adsorbed ions into dissolved forms. Heavy metals incorporated into soils take different chemical forms including (1) dissolved ionic form, (2) electrostatically adsorbed form, and (3) surface complexed form (Darmawan and Wada, 1999). Layer silicate minerals having excess surface negative charge arising from isomorphous substitution adsorb heavy metal cations electrostatically. Due to the nonspecific nature of the coulombic force, the adsorbed heavy metal cations are easily exchanged by other cations. The selectivity coefficient, for example, for Cu –Na exchange on a montmorillonite is near unity (Sposito et al., 1981). On the other hand, oxides and hydroxides of iron and aluminum as well as humic substances bind heavy metal cations very strongly, forming surface complexes, in which the bonding between heavy metal cation and surface functional groups bears some covalency. The surface complexed heavy metal cations are, therefore, not exchanged by common ions in the pore water of soils. Thus, it is expected that the feasibility of the electrokinetic remediation method depends strongly on min-
eralogical composition as well as on soil organic matter content. Studies by Puppala et al. (1997), Reddy et al. (1997), and Grundl and Reese (1997) suggested that the electrokinetic method is not necessarily effective for soils with high adsorption capacity and also for those containing calcium carbonate. To date, however, many researchers have used only artificially polluted model soils made up of pure clays such as kaolinite (Acar and Alshawabkeh, 1996; Yeung et al., 1996; Dzenitis, 1997) and smectite (Reddy et al., 1997; Grundl and Reese, 1997) or their mixture with quartz (Kawachi and Kubo, 1999) in developing and evaluating the electrokinetic remediation technology. There seems to be no systematic experimental study in which the electrokinetic remediation method is applied to a series of natural soils that differ in mineralogical composition. In the present study, therefore, six soils that differ in clay mineralogy and organic matter content were contaminated with salts of copper, lead, and zinc, subjected to electrokinetic remediation treatment, and the behaviors of the heavy metal cations were analyzed through successive extraction.
2. Materials Six soil samples from different locations were used in the present study. They were air-dried and passed through a 2-mm sieve. Some chemical and mineralogical properties determined by standard procedures (Page et al., 1982; Klute, 1986) are listed in Table 1.
Table 1 Selected properties of soil samples Soil name
pH
Organic carbon (g kg 1)
Clay (g kg 1)
Allophane and/ or imogolite (g kg 1)
DCBa Fe2O3 (g kg 1)
Major cation exchanger
Nakajo Harumachi
6.8 5.8
2.98 15.29
382 246
nmb nm
23.8 32.4
Fukuchiyama
6.1
8.81
445
nm
24.9
Pakchong Goshi Choyo
4.9 5.9 6.2
6.18 62.67 22.79
771 385 371
nm 232 304
85.8 54.0 35.5
smectite vermicullite, micaceous, kaolinite micaceous, vermicullite, kaolinite kaolinite, hematite allophane, Al – humus imogolite, allophane
a b
Dithionite – citrate – bicarbonate. No measurement.
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As shown in Table 1, the Nakajo, Fukuchiyama, Harumachi, and Pakchong soils are predominated by crystalline clay minerals. Among these four soil samples, effective CEC of the Pakchong soil is the lowest irrespective of its extremely high clay content, reflecting the predominance of kaolinite in its clay fraction. The Goshi and Choyo soils are similar in that the clay consists exclusively of allophane and imogolite but the Goshi soil has much higher organic matter content. Table 1 also shows that the amounts of Fe and Mn oxides dissolved by dithionite –citrate –bicarbonate (DCB) treatment (Mehra and Jackson, 1960). All the soil samples have fairly high amounts of DCB soluble iron oxide with a maximum of 85.8 g kg 1 for the Pakchong soil, but the amounts of DCB soluble Mn oxides are much lower. Electron microscopy (photos not presented) showed that the iron oxide in the Pakchong soil exists as fine pseudo-hexagonal hematite particles, whereas no discrete iron oxide particles was detectable in other soils. The soils were loaded with 1000 mg kg 1 of Cu(II) and Zn(II) and 500 mg kg 1 of Pb(II) by mixing the soil samples with the calculated amounts of CuSO45H2O, ZnSO47H2O and Pb(NO3)2 and aged for 1 year at field moisture contents. These three metal cations were employed because they are expected to behave differently in soils (Darmawan and Wada, 1999), i.e., Cu(II) and Pb(II) are highly specifically adsorbed by oxides and humic substances but Pb(II) tends to form sparingly soluble sulfates, and Zn(II) is preferred rather by layer silicate minerals.
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The detailed procedure of the sample preparation is given in Darmawan and Wada (1999). Since Cu(II) and Zn(II) were added as soluble sulfate, sulfate ion from these salts would have reacted with Pb(II) ion from Pb(NO3)2 to form sparingly soluble PbSO4 during aging. Although this makes the system complicated, the present combination of salts was selected because PbSO4 is one of the major forms of Pb(II) in heavily polluted soils.
3. Methods The apparatus was made from an open plastic box with dimensions of 220 55 55 mm by partitioning it into four compartments with nylon cloth and attaching an inlet for anolyte, as shown in Fig. 1. The polluted soil samples were packed into the largest central compartment and a pair of graphite electrodes was placed in the two compartments at the two ends. The weight of the packed soil was 320 g for the Nakajo soil, 355 g for the Fukuchiyama soil, 350 g for the Harumachi soil, 335 g for the Pakchong soil, 317 g for the Goshi soil, and 201 g for the Choyo soil. The compartment between the soil and electrode was packed with H-saturated cation exchange resin (Dowex 50W 8) to neutralize OH ion generated at the cathode and to trap heavy metal cations (Wada and Ryu, 1999). The inlet of the apparatus, packed with the soil and resin, was connected to a constant head device containing 10 mM NaCl solution to saturate the
Fig. 1. Schematic illustration of the experimental cell.
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interstitial pores of the soil. After saturation, the level of the anolyte was set 5 mm below the soil surface (Wada et al., 1999) and 20-V potential was applied with a stabilized power supply. Throughout the treatment, the electric current was monitored with an ammeter inserted in the circuit (Fig. 1). After 570 h, the apparatus was disconnected from the power supply and the soil column was cut into six slices of equal length, and air-dried. The soil pH was measured at a soil/water ratio = 1:2.5. Portions (2 g) of the air-dried samples were placed in polycarbonate centrifuge tubes and extracted sequentially with water, 1 M MgCl2 and 6 M HCl. Water and MgCl2 extractions were carried out with 20 ml of water and 16 ml of an MgCl2 solution by shaking for 1 h at room temperature. HCl extraction was performed by digesting with 10 ml of 30% H2O2 to dryness followed by refluxing with 16 ml of 6 M HCl (Asami and Kato, 1977) at boiling point. The extracted Cu(II), Pb(II) and Zn(II) were determined by atomic absorption spectroscopy (AAS) following the procedures described by Amacher (1996) and Reed and Martens (1996). The resin was collected quantitatively and packed in a column. The adsorbed metal cations were extracted by leaching with 1 M NH4NO3 and determined by AAS. All the measurements were carried out in duplicate and the results were averaged.
4. Results and discussion The measured electric currents are plotted against elapsed time in Fig. 2. In general, the electric current decreased rapidly in the first 50-h period followed by gradual decrease in the subsequent 150 h for all soils. The mode of variation and the magnitude of the current were, however, significantly different among soils. For the Nakajo, Fukuchiyama and Harumachi soils, the initial electric current was about 13 –19 mA and it monotonically decreased to the final range of 1.6– 2.7 mA. The trend was quite similar for the Goshi and Choyo soils but both the initial and final values were much lower, i.e., 5.5– 6.2 mA and 0.2 – 0.5 mA, respectively. For the Pakchong soil, on the other hand, the electric current
Fig. 2. Time course of electric current.
increased up to 28 mA after 7 h and rapidly dropped down to < 1 mA after 200 h. The higher initial values of the electric current are obviously due to the higher electrolyte concentration in the pore water. Since indigenous soluble salts are negligible in all soils and the amount of added heavy metal salts were the same, the cause of the lower initial current for humic –allophanic Goshi and allophanic Choyo soils would be that the cations and anions of the added salts were adsorbed on the soil materials to a larger extent (Wada, 1984). The cause of the rapid increase and immediate decrease is not clear but the higher initial electric current in the Pakchong soil supports the contention that the initial contents of water soluble Cu(II), Pb(II) and Zn(II) were the highest in the Pakchong soil (Figs. 3 – 5). The pH profiles along the soil column after the treatment as well as the initial pH values are presented in Fig. 6. Since H + ion was generated by the electrolysis reaction of water continuously at the anode and it migrated toward the cathode, soil pH significantly dropped, particularly in the section neighboring the anode. In the Nakajo, Fukuchiyama, Harumachi and Pakchong soils predominated by crystalline clay minerals, the pH of the soil section next to the anode was below 3. Among these four soils, the Pakchong soil was conspicuous and the pH was far below 3 in the first through fifth sections. In the Nakajo, Fukuchiyama, and Harumachi soils, on the other hand, the magnitude of the pH drop decreased toward the cathode and the pH value of the soil section neighboring the cathode was quite near
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287
Fig. 3. Distribution of three forms of Cu(II) in soils after treatment.
the initial one. The magnitude of pH drops was much smaller in the humic – allophanic Goshi and allophanic Choyo soils and the pH was mostly kept above 5 except in the section neighboring the anode. The plots in Fig. 2 were interpolated by a cubic spline function and integrated over the treatment time to calculate the amount of the electric charge transported through the soil samples. The calculated amount of charge was 7.95 kC for the Nakajo, 5.72 kC for Fukuchiyama, 4.90 kC for Harumachi, 4.76 kC for Pakchong, 1.98 kC for Goshi, and 0.649 kC for Choyo soils. If it is assumed that the electrolytic generation of H + ion is proportional to the electric current, the H + ion supplies in the Goshi and Choyo soils were estimated to be much smaller than that in other soils. However, the pH profiles for the Goshi and Choyo soils are quite similar irrespective of the difference in the amount of the transported charge, and the differences in H + ion concentration estimated from the pH values are much smaller than those expected from the differences in the charge transport
(Fig. 2). These data suggest that the lower electric currents in the Goshi and Choyo soils are not a major cause for their higher pH values. Alternatively and probably the most important cause for that would be the higher acid-buffering capacity of the soils. Allophane and imogolite present in these soils contain large amounts of surface aluminol groups (Wada, 1984). On addition of an acid, HA, the aluminol groups uptake H + ion and the resulting positively charged sites retain the accompanying anion: Al-OH þ Hþ þ A ¼ Al-OHþ 2A :
ð1Þ
With this reaction, the H + ion generated at the anode would have largely been consumed, and in turn, the pH of the interstitial water dropped to a lesser extent. It seems contradictory that the Pakchong soil containing the largest amount of free iron oxide, which can uptake H + ions and anions, exhibited the lowest pH. However, iron oxide in the Pakchong soil was mostly crystalline hematite, whereas it was non-
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Fig. 4. Distribution of three forms of Pb(II) in soils after treatment.
crystalline in other soils. Higher crystallinity of free iron oxide and the predominance of kaolinite with low cation adsorbing capacity would have favored the acidification in the Pakchong soil. The contents of three forms of Cu(II), Pb(II), and Zn(II) in soil slices after electrokinetic treatment are presented in Figs. 3 – 5, together with their initial concentrations. Comparison of these figures indicates that the distribution pattern of Cu(II) and Zn(II) are similar for all soils while that of Pb(II) differed significantly. The distribution of Pb(II) was distinctly different in that the removal rate was significantly lower and the proportion of the MgCl2-extractable fraction was comparable to or even higher than the HCl-extractable fraction (Fig. 4). In the present extraction procedure, water-extractable Cu(II), Pb(II) and Zn(II) would come mostly from soluble salts of these heavy metals. MgCl2extractable fraction would basically represent those retained on layer silicate clay minerals as exchangeable cations. HCl-extractable fraction would corre-
spond to those bound as surface complexes on iron oxides, allophane and humic substances (Darmawan and Wada, 1999) as well as hydroxide precipitates. In addition, for Pb, the MgCl2-extractable fraction would include Pb(II) precipitated as PbSO4 (anglesite) in addition to exchangeable Pb(II), due to the increased solubility of PbSO4 in a concentrated Mg Cl2 solution. This occurs because the activity of sulfate ion is reduced by the formation of soluble MgSO4 complex. The ion speciation based on Pitzer’s model (Gueddari et al., 1983) showed that the ion activity product of (Pb2 + )(SO42 ) would never exceed 10 8 during MgCl2 extraction. Since the solubility product of PbSO4 is 1.62 10 7 (Lindsay, 1979), the precipitate of PbSO4 is expected to have dissolved during extraction and the MgCl2-extractable fraction can be regarded as a sum of exchangeable Pb and PbSO4. Before treatment, copper was found mostly in the strong acid-extractable fraction, except in the Pakchong soil. After the treatment, Cu(II) content re-
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289
Fig. 5. Distribution of three forms of Zn(II) in soils after treatment.
duced markedly in soil sections near the anode to less than 100 mg kg 1 and accumulated in sections near the cathode in the Nakajo, Fukuchiyama, and Harumachi soils. Comparison of Figs. 3 and 6 suggests that the amount of Cu(II) left in the soil is
Fig. 6. Profile of soil pH after treatment.
inversely correlated with pH. The stability of surface complexes of Cu(II) on oxide and humic substances decreases steeply below pH 4 (McKenzie, 1980; Kendorf and Schnitzer, 1980). The distribution pattern of Cu(II) shown in Fig. 3 suggests that Cu(II) was released from surface functional groups, migrated toward the cathode and hydrolyzed to precipitate when it entered into a region of higher pH, as repeatedly reported by many researchers (Yeung et al., 1996; Puppala et al., 1997; Viadero et al., 1998). A noticeable feature in Fig. 3 is that significant amount of MgCl2-extractable, i.e., exchangeable, Cu(II) appeared in the fourth to sixth sections from the anode in the Nakajo, Fukuchiyama, and Harumachi soils. Darmawan and Wada (1999) showed that heavy metal cations loaded to soils are rapidly adsorbed at permanent negative charge on layer silicate minerals, and then slowly transferred to oxides and humic substances to form surface complexes in 50 days. This and the results presented in Fig. 3 suggest that the role of cation exchange sites on
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layer silicate minerals becomes more important in a nonequilibrium migration process, particularly at low pH where the stability of surface complexes is low. Thus, longer treatment time and more electrical power are required for remediation of soils with high cation exchange capacity (Puppala et al., 1997). Fortunately, the cation exchange sites of layer silicate minerals do not show high preference for heavy metal cations (Sposito et al., 1981). The efficiency of electrokinetic remediation may be improved by adding salts of noncontaminant cations, e.g., CaCl2, that compete with heavy metal cations for the exchange site. In the Pakchong soil, the Cu(II) removal rate was the highest among the soils. Total content of the three forms of Cu(II) was below 100 mg kg 1 in the first and second sections and it was about 200 mg kg 1 in the fifth section, i.e., the removal rate was > 80%. The relatively high removal rate would have resulted from the relatively low pH of the original soil and low cation exchange capacity (Table 1) and higher crystallinity, and therefore, lower reactivity of iron oxide minerals. In contrast, Cu(II) was not removed at all from the Goshi soil (Fig. 3). It is apparent that a major cause of this is the high acid-buffering capacity of the soil (Fig. 6). In addition, the extremely high selectivity of humic substances for Cu(II) would have contributed the result. A significant reduction of Cu(II) content in the section neighboring the anode in the allophanic Choyo soil that showed the same pH value to the corresponding section of the humic –allophanic Goshi soil (Fig. 6) supports this view. Fig. 4 shows that the behavior of Zn(II) is basically similar to that of Cu(II) but Zn(II) is much more mobile under the applied electric field. The behavior of Pb(II) shown in Fig. 5 is distinctly different from that of Cu(II) and Zn(II). First, the proportion of MgCl2-extractable fraction is apparently larger in soils dominated by layer silicate minerals before treatment. This must be due to the presence of PbSO4 and its dissolution in 1 M MgCl2 solution as already discussed. Second, the removal rate was significantly lower than that for Cu(II) and Zn(II) and half or more of the remaining Pb(II) was in an MgCl2-extractable form in all soils, except for humic – allophanic and allophanic soils.
Since PbSO4 precipitated in soils does not move either via electrophoresis or via electroosmosis, lower removal rates were expected in soils in which a large proportion of Pb(II) was in a form of PbSO4. Actually the results obtained for the Nakajo, Fukuchiyama, Harumachi and Pakchong soils (Fig. 5) follow the expected trend. In these soils the HCl-extractable fraction, i.e., surface complexed Pb(II), reduced significantly in sections near the anode and seems to have accumulated in the following sections. In contrast, most of the Pb(II) was found in HClextractable fraction and the proportion of MgCl2extractable fractions were low in the Goshi and Choyo soils both before and after the treatment (Fig. 5). This indicates that the precipitation of PbSO4 was greatly suppressed in these soils. Possible mechanism of this is the simultaneous SO42 and Pb2 + adsorption. As already discussed, Pb2 + forms stable surface complexes with surface hydroxyl groups on minerals and carboxyl groups on humic substance. The reaction can be expressed, for example, 2ðSOHÞ þ Pb2þ ¼ ðSOÞ2 Pb þ 2Hþ ;
ð2Þ
where SOH stands for surface hydroxyl groups and carboxyl groups. In acidic condition, some surface hydroxyl groups, typically AlOH and FeOH groups, are protonated and adsorb anions. Among anions, oxo-anions are adsorbed more strongly. The reaction can be expressed as: 2ðAlOHÞ þ 2Hþ þ SO2 4 ¼ ðAlOH2 Þ2 SO4 :
ð3Þ
Since allophanic soils have large amount of the both types of functional groups, the overall reaction is: 2ðSOHÞ þ 2ðAlOHÞ þ Pb2þ þ SO2 4 ¼ ðSOÞ2 Pb þ ðAlOH2 Þ2 SO4 :
ð4Þ
With the simultaneous adsorption of Pb2 + and SO42 , the activity of these ions in solution would be maintained low enough to prevent formation of PbSO4(s). In the present study, H-saturated cation exchange resin was placed between cathode and soil to prevent alkalinization of soil and to trap the migrated heavy metal cations. Wada and Ryu (1999) have examined
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Table 2 Mass balance of Cu, Pb and Zn Heavy metal
Soil name
Initial (g)
Remained in soil
Adsorbed in resin
(g)
(%)
(g)
(%)
Total detected (%)
Cu
Nakajo Harumachi Fukuchiyama Pakchong Goshi Choyo
0.29 0.336 0.326 0.307 0.326 0.181
0.264 0.269 0.245 0.162 0.313 0.170
91.3 80.0 75.2 52.9 96.1 94.3
0.026 0.050 0.070 0.146 0.001 0.001
9.2 14.9 21.5 47.6 0.4 0.6
100.5 95.0 96.8 100.5 96.5 94.9
Pb
Nakajo Harumachi Fukuchiyama Pakchong Goshi Choyo
0.142 0.172 0.170 0.166 0.166 0.095
0.138 0.163 0.149 0.160 0.157 0.092
96.5 94.6 88.0 96.6 94.4 96.7
0.004 0.005 0.007 0.009 0.001 0.000
2.8 3.1 4.1 5.8 0.1 0.2
99.4 97.8 92.2 102.5 94.6 96.9
Zn
Nakajo Harumachi Fukuchiyama Pakchong Goshi Choyo
0.313 0.364 0.347 0.279 0.315 0.173
0.227 0.169 0.091 0.047 0.282 0.173
72.4 46.4 26.1 17.1 89.5 99.8
0.077 0.174 0.217 0.213 0.004 0.005
24.8 47.9 62.6 76.6 1.4 3.0
97.3 94.3 88.7 93.8 91.0 102.8
this in Cu(II) removal from an artificially contaminated soil and found that the alkalinization of the soil neighboring the anode was successfully suppressed and the whole soil was acidified. The pH profile in Fig. 6, however, shows that the alkalinization was prevented but the acid front did not reach the cathodic end of the soil. Therefore, the migrated heavy metal cations hydrolyzed and accumulated in the soil section nearest to the cathode. The reason for the difference between the present result and those by Wada and Ryu (1999) is not clear. One possible reason is that there was no drainage of catholyte in the present study. The mass balances of Cu(II), Zn(II) and Pb(II) are listed in Table 2. The percentages of the total detected heavy metals were mostly >90 except for Zn in the Fukuchiyama soil. Since there was no noticeable interference in the AAS determination, the missing portions of the heavy metals would have associated with the nylon cloth used for separating the resin and soils. Although there are some uncertainties, Table 2 clearly shows that the percentages of Cu(II) and Pb(II) trapped in the cation exchange resin were generally low for all the soil samples than those of Zn(II). Since the objective of the present study was to examine the effect of clay mineralogy on the efficiency of elec-
trokinetic remediation technology, any additional enhancement was not performed. With some enhancement techniques including addition of salts of noncontaminant cations, soil acidification, and drainage of catholyte, the removal rate would be improved, at least, for nonallophanic soils. The results for the Goshi and Choyo soils suggest that the electrokinetic technology would not work for humic – allophanic and allophanic soils, at least for Cu(II) removal. Nevertheless, this may not be a negative finding. The extremely low mobility of heavy metal cations in these soils even under electrokinetic treatment strongly indicates that heavy metal cations incorporated in these types of soils are expected not to be leached into the ground water under the conditions normally encountered in nature. This indicates the possible use of these types of soils for low-cost permeable barrier at waste landfill site.
5. Conclusions . Cu(II) and Zn(II) electrically migrated from the anode toward the cathode to a significant extent in
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soils dominated by crystalline clay minerals. But a lesser drop in pH of soil section near the cathode has led to the precipitation of Cu(II) and Zn(II) into hydroxides. Among these soils, the electrokinetic removal of Cu(II) and Zn(II) in the kaolinitic Pakchong was the highest due to low initial pH, low cation exchange capacity, and lower reactivity of iron oxide minerals of this soil. . The removal rates of Cu(II) and Zn(II) from soils dominated by crystalline clay minerals would be increased by some enhancement techniques, including addition of salts of competing noncontaminant cations, soil acidification, and catholyte drainage. . The high pH-buffering capacities of humic – allophanic and allophanic soils are the major factors that have hindered the electromigration of Cu(II) and Zn(II) in these soils. . Precipitation as sparingly soluble PbSO4 and stable surface complexation with hydroxyl groups of minerals and carboxyl groups of humic substances have resulted in extremely low migration of Pb(II) from all soils under study. . Strong retention of heavy metal cations by humic – allophanic and allophanic soils and their extremely low mobility even under electrical field indicate the possible use of these soils as an ideal permeable barrier at landfill site.
Acknowledgements We thank Dr. Mochizuki of Experimental Farm of Kyushu University for providing a soil sample. This study was supported in part by a Grant-in-Aid for Scientific Research (#11660066) from the Japanese Society for Promotion of Sciences.
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