Immobilization of Heavy Metals by Polynuclear Aluminium and

Therefore, the aluminium based binding agents may be used for the gentle remediation of soils polluted by nickel, copper, zinc, or cadmium. Introducti...
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Environ. Sci. Technol. 1997, 31, 1452-1462

Immobilization of Heavy Metals by Polynuclear Aluminium and Montmorillonite Compounds BARBARA LOTHENBACH,* GERHARD FURRER, AND RAINER SCHULIN Institute of Terrestrial Ecology, ETH Zurich, Grabenstrasse 3, CH-8952 Schlieren, Switzerland

Modified montmorillonite compounds and polynuclear Al13 were investigated as potential binding agents to reduce heavy metal solubility in soil solutions. The experiments were carried out in batch reactors with solutions/suspensions containing montmorillonite, polynuclear Al13, Al-montmorillonite, or Al13-montmorillonite. Nickel, copper, and zinc were adsorbed specifically on Al-montmorillonite and Al13montmorillonite. These three metals were also incorporated in the interlayers of Al-montmorillonite and Al13-montmorillonite during the aging processes. Al-montmorillonite was most effective in the pH ranges 6-8 for nickel and zinc, 4-6 for copper, and 7-9 for cadmium. The immobilization of the heavy metal cations by Al13-montmorillonite was restricted to a smaller pH range. The polynuclear Al13 decreased the dissolved concentrations of nickel, copper, and zinc by specific binding from pH 6.1 to pH 7.5. In the pH range 5.86.1 nickel, copper, zinc, cadmium, and lead formed soluble complexes with Al13, which remained dissolved during 30 weeks. In summary, the four binding agents were found to immobilize nickel, copper, zinc, and cadmium, whereas the effect on the solubility of lead was rather small. Therefore, the aluminium based binding agents may be used for the gentle remediation of soils polluted by nickel, copper, zinc, or cadmium.

Introduction High concentrations of heavy metals in soils may cause longterm risks to ecosystems and humans. The accumulation of heavy metals is a result of the excessive application of sewage sludge and fertilizers or atmospheric deposition (1, 2). In calcareous soils, heavy metal solubility is small, whereas in neutral or acidic soils a relevant fraction of the metals can dissolve and become available for uptake by plants. The remobilization of heavy metals as a consequence of acidification can also represent a major threat to groundwaters (3). Rapid soil acidification can be caused by changes in land use practices or, more slowly, by acidic atmospheric deposition. In general, the uptake of heavy metals by plants leads to increased heavy metal contents in food chains. The bioavailability of heavy metals in soils is mainly determined by organic matter content, mount of iron and manganese oxides, clay minerals, and pH regime. Various methods have been developed for the remediation of heavy metal contaminated soils (4-7). Traditional soil remediation methods, such as disposal of contaminated material or soil washing, are destructive and in general also * Corresponding author telephone: +41 1 633 60 12; fax: +41 1 633 11 22; e-mail address: [email protected].

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rather expensive. There is a necessity for gentle remediation methods that do not destroy soil structure or fertility and are cheap and suitable for application on arable land. In recent years, many attempts have been made to find heavy metal accumulating plants that are particularly able to extract metals from soils (e.g., ref 6). Under field conditions, most of these plant species exhibit limited growth or insufficient transfer coefficients for heavy metals. The extraction of metals from polluted soils by the present generation of hyperaccumulators lasts decades rather than years. There is therefore still a necessity for abiotic remediation methods. Traditionally, liming of soil has been used to increase the alkalinity, thereby reducing the mobility of metals. However, an increase of pH can cause a number of negative effects such as reduced manganese and phosphate availability to plants (8, 9), changes in soil biocenoses, increased nitrate concentrations in drainage water, or even a mobilization of copper and lead due to the decay of humus (10). This paper presents the results of the testing of binding agents suitable for the immobilization of heavy metals without modification of pH. The investigated compounds immobilize soluble fractions of heavy metals in soil solutions due to adsorption reactions or incorporation into the solids. Binding Agents. Clay minerals are potential binding agents for pollutants as they have a large specific surface area and a high cation exchange capacity. In general, clay minerals exhibit both a pH-dependent and a permanent negative surface charge, which is caused by isomorphous substitution. The permanent charge is for the most part compensated by cations intercalated in the interstitial space (11). The binding mechanism of cations on the permanent charged sites is mainly characterized by electrostatic interactions. Therefore, metal cations bound in such a way can readily be exchanged by other cations, e.g., by calcium or magnesium (12). The pH-dependent charge is located at the edge sites, where the surface hydroxyl groups can be protonated or deprotonated, depending on the solution pH. The adsorption of heavy metal cations by hydroxyl groups is governed by covalent binding and therefore chemically specific, i.e., equally charged cations can exhibit different affinities for the surface sites. In the case of smectite minerals, the capacity of specific adsorption is smaller than that of cation exchange. However, at high pH and at trace concentrations of heavy metals, specific adsorption is often the dominant binding mechanism (12, 13). The adsorption properties of clay minerals toward heavy metals can be improved by the addition of aluminium (14, 15). After partial neutralization, the aluminium precipitates in the form of partially neutralized hydroxide layers at the mineral surfaces and in the interstitial space. This compensates most of the cation exchange capacity (CEC). The structure and characteristics of the modified clay minerals depend on the preparation conditions (16, 17). As known from binary aluminium hydroxide solids, clay minerals with aluminium hydroxide coatings can specifically adsorb heavy metal cations (14, 15). Specific adsorption sites for heavy metals can also be offered by the polynuclear aluminium complex [AlO4Al12(OH)24(H2O)12]7+ (referred to as Al13 or Al137+, respectively) (18, 19). This polymer is formed in partially neutralized solutions at a minimal Al(III) concentration of 10-5 M and at pH values of 5.4-6.0 (20). Long-term aging experiments have shown that in alkaline solutions the Al13 unit may slowly convert to amorphous Al(OH)3(s) (21). During the structural rearrangement, incorporation of small cations like copper, zinc, or nickel into the solid matrix is likely. This has been

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TABLE 1. Physical Properties of Montmorillonite, Al-Montmorillonite, and Al13-Montmorillonite d(001) (Å)

montmorillonite (SWy-1) Al-montmorillonite Al13-montmorillonite a

3 weeks

20 weeks

50 weeks

BET surface area (m2 g-1)a

CEC (cmolC kg-1)a

12.0 15.9 18.8

13.8 14.6

14.0 14.4

20 60 220

86 14 4

Weight refers to montmorillonite without aluminum or Al13.

observed in the case of amorphous aluminium hydroxide (22, 23). Another binding agent can be formed from the addition of polynuclear Al13 to montmorillonite: pillared montmorillonite (24-26). Aluminium pillared clay minerals are characterized by a large surface area and a high thermal stability and are widely used in refinery processes as catalysts (27). In this study, we investigated the adsorption of nickel, copper, zinc, cadmium, and lead in aqueous suspensions by montmorillonite, Al13, and montmorillonites that were coated and intercalated with aluminium hydroxides (Al-montmorillonite) or with Al13 (Al13-montmorillonite) as potential binding agents for the gentle remediation of soils polluted by heavy metals.

Experimental Section Materials. Montmorillonite originating from Crook County, Wyoming (SWy-1, Clay Minerals Society), was used. The particle fraction e2 µm (Stokes’ diameter) was separated by sedimentation technique. The sodium form of the clay was obtained by washing this fraction three times with 1 M NaCl followed by four washings with Nanopure water. The stock suspension contained 15 g of montmorillonite/L and was stored in the dark. Al13 Stock Solution. A total of 800 mL of 0.2 M AlCl3 was filtered (0.01 µm) and titrated with 612 mL of 0.64 M NaOH at a rate of 3 mL/h until an OH/Al ratio of 2.45 was reached. The resulting clear solution was filtered (0.01 µm). The 27Al-NMR spectra, obtained from a Bruker 400 MHz spectrometer operating at 104.3 MHz, exhibited one single sharp peak at 62.5 ppm, representing the central tetrahedral aluminium in the Al13 polymer (28). The total aluminium concentration determined by ICP-AES was 0.118 M. The Al13 stock solution was checked for turbidity with a laser beam (632.8 nm). This test did not show any Tyndall effect during the whole period of the experiments. Al13-montmorillonite (montmorillonite intercalated with Al13) was prepared by adding 200 mL of water and 68 mL of 9.1 mM Al13 solution [0.118 M Al(III)] to 266 mL of the montmorillonite stock suspension until a ratio of 2 mmol of Al(III)/g of clay was reached. A ratio of 2 mmol of Al(III)/g of clay was chosen as 1 g of montmorillonite adsorbs at most 2 mmol of Al(III) in the form of Al137+ (24, 29). Al-montmorillonite (hydroxy-aluminium montmorillonite) was prepared by adding 216 mL of a 37 mM AlCl3 solution to 266 mL of the montmorillonite stock suspension. Thus, a ratio of 2 mmol of Al(III)/g of clay was reached, the same ratio as in the case of Al13-montmorillonite. The suspension was titrated with 196 mL of 0.1 M NaOH at a rate of 2 mL/h until an OH/Al ratio of 2.45 was reached. After aging 3 days at room temperature, both the Al13montmorillonite and the Al-montmorillonite suspension were washed three times with Nanopure water and used for the experiments within 1 month. All indications of weights are referred to untreated montmorillonite. The presence of aluminium in the interlayer increased the weight by 15% for Al-montmorillonite and by 17% for Al13-montmorillonite. All chemicals were reagent grade.

Characterization of Materials. X-ray diffraction of the air-dried Al-montmorillonite showed an expanded d(001) spacing of 15.9 Å relative to that of Na-montmorillonite of 12.0 Å (Table 1). Harsh and Doner (14) reported for a similarly prepared Al-montmorillonite a spacing of 16.1 Å. The expansion of about 4 Å in comparison to Na-montmorillonite indicated that aluminium hydroxides are precipitated in the interlayer. The absence of higher order (00l) reflections indicated that the aluminium hydroxides are irregularly distributed in the interstitial space. The X-ray diffraction of Al13-montmorillonite showed an expanded d(001) spacing of 18.8 Å followed by several higher order (00l) reflections (9.5, 6.3, 4.7, 3.7, 3.2, and 2.7 Å). These indicated a uniform distribution of polynuclear hydroxyaluminium cations in the interlayer. Similar results were obtained by Bukka et al. (26) and Michot et al. (30). The external BET surface area of the clay samples was determined by adsorption of N2 at 77 K and was found to increase considerably in the presence of aluminium (Table 1), due to the expansion of the interlayer spacing. Cation exchange capacity (CEC) of montmorillonite, Almontmorillonite and Al13-montmorillonite, was obtained by displacing the exchangeable cations by BaCl2, a procedure described by Hendershot and Duquette (31). Approximately 0.3 g of clay mineral was suspended in 30 mL of 0.1 M BaCl2 solution and shaken for 1 week. After centrifugation, the clear supernatant was removed, and the procedure was repeated three times with 0.025 M BaCl2 and finally with 0.01 M MgSO4. The supernatants were analyzed for Mg2+ by flame atomic adsorption spectroscopy (AAS) on a Varian AA400 spectrophotometer. CEC was calculated from the difference of added and measured Mg2+. The CEC of originally 86 cmolC/ kg montmorillonite dropped drastically after the clay mineral had been coated with aluminium or Al13 (Table 1). Solid-state 27Al-NMR spectra were recorded on a Bruker AMX 400 operating at 104.3 MHz using magic angle spinning (MAS) and referenced externally to dissolved Al(H2O)63+ (Figure 1). The 27Al-NMR spectra of the montmorillonite revealed a signal at -4.1 ppm, which was assigned to octahedral aluminium, and a smaller signal at 61.5 ppm, which was assigned to tetrahedrally co-ordinated aluminium in the montmorillonite lattice (from isomorphous substitution). Four side bands of the octahedral aluminium were clearly visible. The spectra of freshly prepared Al13-montmorillonite and Al-montmorillonite showed signals at -1.8 and 57.3 ppm and at -1.2 and 58.0 ppm, respectively. The signals at 57.3 and at 58.0 ppm could be assigned to the central tetrahedral aluminium in the Al13 polymers in the interlayer (32). As the signal at 58.0 ppm was markedly smaller than the corresponding one observed for Al13-montmorillonite only a small amount of Al13 polymers were probably present in the interlayer space. Methods. Batch experiments were performed in glass vessels and with a solution or suspension volume of 100 mL. Water, NaClO4, montmorillonite (untreated or previously treated with Al13 or Al, respectively), the chloride salts of heavy metals, and HCl or NaOH were added to achieve a final montmorillonite concentration of 2 g/L, an ionic strength of 0.1, and a heavy metal concentration of 100 µM. For the first experiments with cadmium, NaClO4 was replaced by NaCl

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TABLE 2. Cation Radii, Rate of Water Exchange, pK1 and pKSO Values of Metal Cations cation radiia (octahedral water pK1, co-ordination) exchange I ) 0.1, (pm) rateb (s-1) T ) 25 °C Ni2+ Cu2+ Zn2+ Cd2+ Pb2+ Al3+ a

83 87 88 109 133 67.5

From ref 38.

b

3 × 104 1 × 109 7 × 107 3 × 108 7 × 109

10.2c 7.7c 9.2c 10.3d 7.8c 5.4c

From ref 39. c From ref 36.

d

solid phase

pKSO, I ) 0.1, T ) 25 °C

Ni(OH)2 CuO ZnO CdCO3 PbCO3

-16.5e -19.7d -16.2d -12.9d -12.0c

From ref 35. e From ref

37.

FIGURE 1. 27Al-NMR spectra of Al-montmorillonite, Al13-montmorillonite, and montmorillonite. x: side bands. and a heavy metal concentration of 1000 µM was used. The batch vessels were purged with argon to reduce CO2 partial pressure, closed tightly, and shaken at 200 rpm at room temperature during 4 weeks. The pH was measured under argon several times and, when necessary, adjusted with acid or base. After 4 weeks, the pH was measured again, and 10 mL of the suspensions was taken and centrifuged at 5000 rpm (3990g) for 30 min. The supernatants were filtered (cellulose nitrate, 0.01 µm), and the concentrations of heavy metals in the filtrates were analyzed by flame AAS. To quantify the exchangeable fraction of the adsorbed heavy metals, 9 mL of 37 mM Ba(ClO4)2 solution was added to the remaining 1 mL of wet sediment, thus the suspension contained 33 mM Ba(ClO4)2. The samples were shaken for 2 h and analyzed for dissolved heavy metal concentrations. The remaining suspensions were stored in the dark, and pH and heavy metal concentrations were analyzed again after 30 and 60 weeks. The measurements of pH were carried out under argon with a combined pH electrode (Metrohm 6.0204.100) connected to a digital voltmeter (Metrohm 713). This equipment was calibrated prior to the measurements by acid-base titration in the respective ionic medium. Sorption Modeling. The program GRFIT (33) was used for fitting the adsorption constants. GRFIT uses a non-linear, least-squares optimization technique to calculate equilibrium constants from chemical data. The speciation of the heavy metals was calculated using the program MICROQL (34) with thermodynamic constants from Smith and Martell (35, 36) and Baes and Mesmer (37), corrected for ionic strength using the Davies equation. Values for the formation of the thermodynamically favored solid phases Ni(OH)2, CuO, ZnO, CdCO3, and PbCO3 are compiled in Table 2. PbCO3 and CdCO3 phases were taken into account as the batch vessels were not completely gas tight. The observed precipitation of CdCO3 was slow and was modeled employing a partial CO2 pressure of 9 × 10-9 atm (9 × 10-4 Pa) after 4 weeks and 6 × 10-6 atm (6 × 10-1 Pa) after 30 weeks. Different sorption models were used to fit the results gained in the batch experiments. A variety of surface models for montmorillonite can be found in the literature (e.g., refs 40

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and 41). The adsorption on SWy-1 montmorillonite was modeled using the non-electrostatic model of Bradbury and Baeyens (42). The model equations are given in Table 3. Besides fixed charge exchange sites (X-) and edge hydroxyl sites with weak affinity (SwOH), this model assumes the existence of edge hydroxyl sites with high affinity for heavy metals (SsOH). The high affinity sites have a relatively low capacity (Table 4). Because these constants could not be determined with our experimental data, the values given by Bradbury and Baeyens (42) were used. Deprotonation of the Al137+ polymer has been described by Ludwig (43) and by Furrer et al. (44). Heavy metal adsorption by Al13 has been fitted using the non-electrostatic deprotonation model that was described by Ludwig (43) for an Al13 concentration of 0.5 mM. Heavy metals adsorb on the deprotonated Al13+ releasing a further proton (Table 3). For Al-montmorillonite and Al13-montmorillonite, no appropriate sorption models are available. For Al13-montmorillonite, it was assumed that the Al13 in the interlayer exhibited the same acid-base behavior as the dissolved Al13 polymers. The cation exchange constants for lead and cadmium were fitted. For nickel, copper, and zinc, cation exchange constants obtained for the adsorption on montmorillonite were used. The adsorption of heavy metals on the surface sites of Al-montmorillonite was fitted using the triple-layer model (TLM) of Yates et al. (45) and Davis and Leckie (46) with the constants and surface parameters of aluminium-modified silica (47). H+, OH-, and other strongly binding ions such as heavy metals are incorporated in inner-sphere surface complexes located in the innermost layer (48, 49). We assumed that the aluminium hydroxide covered most of the montmorillonite surface. This leads to new adsorption sites (AlOH). For the remaining cation exchange sites on Almontmorillonite, the exchange constants previously fitted for montmorillonite were used. Only for lead could a new cation exchange constant be determined. We also modeled complexes that included several heavy metals (e.g., Al13Ni2) or several sorption sites [e.g., (AlO)2Ni], but as the calculated sum of errors in the square of all data points, χ2, was larger than for the 1:1 complexes; 1:1 complexes were assumed.

Results A clear change in dissolved cadmium concentrations was caused when Al(III) was present in the montmorillonite suspensions (Figure 2 and Table 5). At pH values above 6, dissolved cadmium concentrations decreased in the presence of aluminium, whereas at lower pH values aluminium did not enhance cadmium adsorption. The effect was clearly dependent on the amount of aluminium added. Adsorption of Heavy Metals. In Figure 3, panels a-e, the influence of montmorillonite, Al-montmorillonite, Al13montmorillonite, and Al13 on dissolved heavy metal concen-

TABLE 3. Surface Sites Reactionsa and Constants montmorillonite

log Kb

Al-montmorillonite

log Kf

2XNa + Me2+ T X2Me + 2Na+ SsOH + H+ T SsOH2+ SsOH T SsO- + H+ SsOH + Me2+ T SsOMe+ + H+ SwOH + H+ T SwOH2+ SwOH T SwO- + H+ SwOH + Me2+ T SwOMe+ + H+

log KXc 4.6 -7.9 1.6d 4.6 -7.9 log KSwc

2XNa + Me2+ T X2Me + 2Na+ AlOH + H+ T AlOH2+ AlOH T AlOMe+ + H+

log KXc 8.0 -10.2 log KAlc

Al13

log Ke

Al13-montmorillonite

log Ke

Al137+ T Al135+ + 2H+ Al137+ T Al133+ + 4H+ Al137+ T Al132+ + 5H+ Al137+ T Al13+ + 6H+ Al137+ Me2+ T Al13Me2+ + 7H+

-11.1 -22.8 -28.5 -35.0 log KAl13c

2XNa + Me2+ T X2Me + 2Na+ Al137+ T Al135+ + 2H+ Al137+ T Al133+ + 4H+ Al137+ T Al132+ + 5H+ Al137+ T Al13+ + 6H+ Al137+ Me2+ T Al13Me2+ + 7H+

log KXc -11.1 -22.8 -28.5 -35.0 log KAl13c

a XNa, fixed exchange sites; Me2+, heavy metal; SsOH, edge hydroxyl sites with high affinity; SwOH, edge hydroxyl sites with weak affinity; AlOH, Al-montmorillonite surface site. b Values from ref 42. c Constants that were determined by fitting for the respective heavy metal. d For nickel, a constant of -0.1 was used (42). e Values from ref 43. f Values from ref 47.

TABLE 4. Parameters for Montmorillonite, Al13, Al13-Montmorillonite, and Al-Montmorillonite montmorillonitea CEC (mequiv/g) SsOH (mmol/g) SwOH (mmol/g) AlOH (mmol/g) Al13 (mmol/g) Al13 (mmol/L)

Al13b

0.86 0.002 0.040

Al13-montmorilloniteb

Al-montmorillonitec

0.04

0.14 2

0.154 0.308

a

Values from ref 42. b Values from ref 43. c Values from ref 47; parameter for triple layer model, surface area 60 m2/g; inner capacitance 1.25 F/m2; outer capacitance, 0.20 F/m2.

FIGURE 2. Effect of different amounts of Al(III) on the adsorption of cadmium by montmorillonite in batch experiments. The suspensions contained 0.1 M NaCl, 2 g L-1 montmorillonite, 0, 0.2, 2, or 4 mM AlCl3, and 1 mM CdCl2 and HCl or NaOH added in this sequence. Dissolved cadmium concentrations and pH were measured after 4 weeks. The solid lines have not been calculated; they serve as eye guides. trations is shown. Solutions free of binding agents were used as reference systems. The dotted lines represent the calculated dissolved concentrations of the thermodynamically favored solids Ni(OH)2, CuO, ZnO, CdCO3, and PbCO3, respectively. The bold lines represent the fitted and calculated dissolved concentrations of the five metals in the presence of montmorillonite as affected by cation exchange and specific surface complexation. At high pH values, dissolved heavy metal concentrations in the presence of montmorillonite are limited by precipitation. With decreasing pH, the dissolved metal concentrations increase to approximately 80%. Compared to pure montmorillonite, the presence of Almontmorillonite or Al13-montmorillonite led to a decrease of dissolved heavy metal concentrations. This effect was very

pronounced for zinc, nickel, and copper; less pronounced for cadmium; and almost insignificant for lead. Zinc and copper adsorbed better to Al-montmorillonite than to Al13montmorillonite, whereas for nickel, lead, and cadmium, the adsorption on the two modified clay minerals was similar. The dashed lines represent the fitted and calculated heavy metal concentrations in the presence of Al13- and Almontmorillonite, respectively. At lower pH values, less heavy metal cations were adsorbed to Al13- or Al-montmorillonite than to montmorillonite. The Al13 polymer was effective in immobilizing nickel, zinc, and cadmium, whereas the effect on the dissolved concentrations of copper and lead was rather small. The fitted concentrations of the metals in the presence of Al13 are shown by the solid lines in Figure 3, panels a-e. Exchange of Adsorbed Heavy Metals by Barium. Heavy metals adsorbed on montmorillonite were considerably remobilized after the addition of 33 mM Ba(ClO4)2 (Figure 4, panels a-e). Below the pH50 values, the values at which 50% of the heavy metals are adsorbed or precipitated, all heavy metals were nearly fully remobilized by the 330-fold excess of barium cations, indicating that these heavy metals were adsorbed mainly nonspecifically by electrostatic forces. The addition of barium to the Al13 solutions and suspensions did not show any significant remobilization of nickel, copper, zinc, or lead, whereas adsorbed cadmium was partially remobilized by the excess barium (Figure 5, panels a-e). In suspensions with Al13-montmorillonite or Al-montmorillonite virtually no nickel, copper, and zinc was remobilized (Figure 6, panel a-c). In contrast, cadmium adsorbed on Al-montmorillonite and Al13-montmorillonite was readily exchanged, while lead was partially remobilized (Figure 6d,e). Influence of Time. In order to investigate the time effect on heavy metal adsorption, the measurements were carried out after 2, 4, 30, and 60 weeks. Copper was measured after 4 weeks only. In Figures 7 and 8, the results are exemplified

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TABLE 5. Calculated Constants for Montmorillonite, Al13, Al13-Montmorillonite, and Al-Montmorillonitea montmorillonite

nickel copper zinc cadmium lead a

2 weeks 4 weeks 30 weeks 4 weeks 2 weeks 4 weeks 30 weeks 4 weeks 30 weeks 4 weeks 30 weeks

Al13

Al13-montmorillonite

log KSw

log KX

log KAl13

-3.9

2.7

-37.8

-1.9

2.8

-37.0

-3.2

2.7

-36.8

-3.8

2.6

-38.2

-2.1

2.8

-36.8

For definitions of the calculated constants cf. Table 3.

b

log KAl13

log KX

-37.6 -33.7 -29.7 -31.8 -36.8 -37.2 -30.4 -38.7 -38.6 -36.6 -37.1

b b b b b b b 4.6 4.8 5.0 4.6

Al-montmorillonite log KAl

log KX

-0.1 2.0 3.1

b b b

1.9 2.9 -2.0 -4.2 -0.8 -0.9

b b b b 4.3 4.0

The constants KX calculated for montmorillonite were used.

FIGURE 3. Effect of montmorillonite, Al-montmorillonite, Al13-montmorillonite, and Al13 on dissolved concentrations of nickel, copper, zinc, cadmium, and lead in batch experiments after 4 weeks. The dotted lines represent the theoretical heavy metal concentrations in the presence of the solids Ni(OH)2, CuO, ZnO, CdCO3, and PbCO3, respectively, based on stability constants from Smith and Martell (35, 36) and Baes and Mesmer (37). The dashed and solid lines represent the fitted metal concentrations in the presence of montmorillonite, Al-montmorillonite, Al13-montmorillonite, or Al13 (for reactions and constants cf. Tables 3 and 5). by nickel and lead, respectively. No aging effect was observed in the absence of minerals (Figures 7a and 8a) or in the presence of montmorillonite (Figures 7b and 8b). The adsorption of nickel on both Al-montmorillonite and Al13-montmorillonite increased clearly during the first 30 weeks (Figure 7, panels c and d). The time effect in the case of Al-montmorillonite is less pronounced than in the case of Al13-montmorillonite. In contrast, the adsorption of lead on the two modified montmorillonite compounds was not affected by aging (Figures 8, panels c and d). Similarly, the dissolved concentrations of nickel and lead in presence of Al13 remained constant during the observed time period (Figures 7e, and 8e). In all five systems presented by Figures 7 and 8, the behavior of zinc and cadmium was very similar to the behavior of nickel and lead, respectively. The exchangeability of these four metals by excess barium did not change with time in any of the investigated systems. Stability of the Al13 Polymers. The polynuclear aluminium complex [AlO4Al12(OH)24(H2O)12]7+ is composed of 12 aluminium octahedra and one central aluminium tetrahedron (50). Each of the octahedra carries one terminal water ligand, which means that Al137+ is a polyprotic acid with 12 identical functional groups. Deprotonation of the terminal water

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ligands results in hydroxyl groups that exhibit excellent binding characteristics, similar to those of aluminium hydroxide surfaces. Deprotonation of the Al137+ species starts around pH 6, depending on the total polymer concentration and leads to aggregation and subsequent precipitation of the polymers (44). Bottero et al. (21) have shown that at OH/Al ratios above 2.6, in solutions with I ≈ 0.5 NaCl and [Al13] ) 8 mM, the Al13 units are transformed partially to amorphous Al(OH)3(s) or bayerite within several days. In our experiments, the transformations to amorphous aluminium hydroxide were observed in solutions with I ) 0.1 NaCl, [Al13] ) 0.3 mM, and 0.1 mM nickel, zinc, cadmium, or lead, respectively. At pH values above 6.9 (OH/Al ratio > 2.9), a whitish gel appeared within a few minutes. At pH values between 5.8 and 6.1 (OH/ Al ratio: 2.70-2.75), the formation of Al13 aggregates, which also incorporate heavy metals, is presumed: filtration (0.01 µm) and centrifugation (cut off Stokes’ diameter ) 0.10 µm) experiments revealed that about half of the total dissolved aluminium is present as particles in the size range of 0.010.10 µm. These aggregates were persistent at least 30 weeks as indicated by the clear solutions and by the absence of a Tyndall effect (laser beam, λ ) 632.8 nm). After 60 weeks, the solutions were turbid. At pH values below 5.1 (OH/Al ratio