Environ. Sci. Technol. 2004, 38, 1522-1528
A Two-Front Leach Model for Cement-Stabilized Heavy Metal Waste M O H A M M A D Z . I S L A M , †,‡ L I O N E L J . J . C A T A L A N , * ,† A N D ERNEST K. YANFUL‡ Department of Chemical Engineering, Lakehead University, 955 Oliver Road, Thunder Bay, Ontario, Canada P7B 5E1, and Department of Civil and Environmental Engineering, The University of Western Ontario, London, Ontario, Canada N6A 5B9
Quantitative scanning electron microscope (SEM) studies of cement-stabilized waste specimens exposed to a leaching solution at constant pH in the range 4-7 have shown that the acid neutralization capacity (ANC) of the waste matrix is consumed at two consecutive leaching fronts. The first front is associated with the dissolution of portlandite (Ca(OH)2) and the partial reaction of calcium silicate hydrate (CSH) gel. The second front marks the dissolution of Ca-Al hydroxy sulfate minerals. The advancement of the first front is limited by the diffusion of OH- ions from the first front toward the leaching solution. The advancement of the second front, however, is controlled by the diffusion of H+ ions from the leaching solution toward the second front. Leaching of copper, zinc, and lead only occurs between the second front and the specimen surface. The leaching behavior of metals is modeled by considering that metals are leached from the waste matrix as a result of the advancement of the second front. The proposed model takes into account the leachable metal fraction in the waste matrix and the effect of metal remineralization on metal mobility.
Introduction Solidification/stabilization (s/s) is the second most common type of source control technology implemented at Superfund remedial sites, and it has been used at more than 160 of these sites to treat contaminated soil, sediment, and sludge since 1982 (1). A majority of these projects have used cementitious binders to treat metal-containing wastes. S/s is also listed as the best demonstrated available technology (BDAT) for over 50 commonly produced hazardous industrial wastes in the U.S. land disposal regulations (2). In cement-based s/s, the mobility of metal contaminants and their likelihood of release to the environment are decreased through mechanisms such as precipitation as metal hydroxides, adsorption or ion substitution in cement hydration products, and physical encapsulation in the solid matrix (3). The effectiveness of s/s is assessed by performing both chemical and physical tests after the solidified material has cured. The most common tests are the toxicity charac* Corresponding author phone: (807)343-8573; fax: (807)343-8928; e-mail:
[email protected]. † Lakehead University. ‡ The University of Western Ontario. 1522
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teristic leaching procedure (TCLP) and the unconfined compressive strength (UCS) measurement (1). Ideally, the potential impact of s/s waste on environmental parameters (e.g. groundwater quality) at disposal sites can be evaluated through a risk assessment process that makes use of data from leach tests and predictive leach models (4). However, further understanding of leaching mechanisms is required before reliable predictions of the long-term releases of contaminants can be obtained using leach models. Several leach models taking into account the transport of contaminant species in pore water and the reaction of cement phases with the leachant have been proposed to predict the leaching behavior of s/s waste. Hinsenveld and Bishop (5) and Baker and Bishop (6) describe a shrinking unreacted core (SUC) leach model that is based on the following hypotheses: (1) the dissolution of the cement matrix and the solubilization of metal ions occur at a well defined front separating a leached shell (in contact with the leachate) from an unreacted specimen core, (2) the acid neutralization capacity (ANC) of s/s waste consists primarily of calcium hydroxide (Ca(OH)2), and (3) the leaching process is limited by the diffusion of acid species (H+) through the leached shell. The first two hypotheses of the SUC model are contradicted by recent SEM observations of s/s waste (7) after leaching with acid or neutral solutions. These observations show that the ANC of cement-based s/s waste is contributed not only by calcium hydroxide but also by the CSH gel and unhydrated tricalcium silicate (Ca3SiO5). Furthermore, two consecutive leaching fronts advancing from the surface toward the interior of the s/s waste were identified. The first front is associated with the dissolution of Ca(OH)2 and partial reaction of CSH gel, whereas the second front results from the dissolution of Ca-Al hydroxy sulfates such as ettringite (Ca6Al2(SO4)3OH12‚26H2O)andmonosulfate([Ca2(Al,Fe)OH6]2(SO4).xH2O). Metal leaching only occurs in the outer layer between the s/s waste surface and the second leaching front. Baker and Bishop (6) found that the SUC model overpredicts heavy metal releases and introduced retardation factors to reconcile model predictions with the experimental data. However, SEM and X-ray energy dispersive spectroscopy (EDX) analyses of leached s/s waste (7) have shown that the presence of a metal remineralization zone behind the second leaching front provides a sink for heavy metals and drastically limits their releases to the leachate. Remineralization occurs when part of the heavy metals dissolved in the outer layer diffuses inward and reprecipitates as hydroxides at higher pH values. Because the SUC model does not consider the effect of metal remineralization, this effect could partially explain the discrepancies between predicted and measured metal releases. Numerical leach models that account for the coupling between mineral dissolution and ion diffusion have also been developed and applied to describe the release of lead and arsenic from s/s waste (8-10). These models assume that the pore water pH is solely controlled by the dissolution of Ca(OH)2 coupled with the mass transfer of hydroxide ions. Recently, the effect of other soluble alkaline salts such as KOH and NaOH on pH was also considered in a modified coupled dissolution/diffusion model (11). However, these models do not account for the sequential dissolution of the CSH gel and Ca-Al hydroxy sulfate minerals and its effect on the leaching processes. Another limitation is the assumption that leachable heavy metals are only present as metal oxides or hydroxides within the pore structure, whereas there is ample evidence that heavy metals can also be 10.1021/es0348400 CCC: $27.50
2004 American Chemical Society Published on Web 01/28/2004
incorporated in cement hydration products either by adsorption or coprecipitation (3, 12, 13). Mainguy et al. (14) developed a coupled dissolution/ diffusion model, named DIFFU-Ca, for predicting the chemical degradation of cement paste and mortar by the leaching action of pure water. This model accounts for the incongruent dissolution of the CSH gel (15) and was adapted by Matte et al. (16) to include the dissolution of cement anhydrous phases. Measured calcium profiles in the solid and calcium fluxes to the leachate were successfully simulated, but this model is not applicable to heavy metal leaching. Park and Batchelor (17) presented a multicomponent numerical leach model (SBLEM) coupled with a chemical speciation code (SOLTEQ-B), which is an extension of MINTEQA2 (18) to cementitious systems. This model considers the diffusion of ionic species and the incongruent dissolution of the CSH gel. It represents an expansion of earlier leach models (19, 20) that assumed simple pH dependent precipitation or adsorption reactions for metals. These earlier models were successful at predicting the phenomenon of heavy metal remineralization. To our knowledge, no model is currently available in the literature to describe the advancement of consecutive leaching fronts associated with the sequential dissolution of various cement hydration products in the s/s waste matrix and the concomitant release of heavy metals. In this paper, we develop such a model. Our analytical two-front leach model also takes into consideration the effect of remineralization on heavy metal release. This is done by adapting the basic equations of the SUC model (5, 6). As a result, the two-front leach model retains the same basic form as the SUC model. The ability of the new model to simulate the experimental leaching data presented in ref 7 is assessed.
FIGURE 1. Reacted ANC versus time.
Experimental Section
FIGURE 2. Ca release versus acid consumption.
Cylindrical specimens of s/s treated waste (7.6-cm diameter and 12.7-cm height) were cast by mixing ordinary Portland cement (OPC) and a synthetic alkaline sludge containing known amounts of heavy metals. The synthetic sludge was prepared by adding 6 M NaOH to a solution of 0.1 M Cu(NO3)2, 0.1 M Zn(NO3)2, and 0.1 M Pb(NO3)2 to an end point pH of 9.0. The weight fraction of wet sludge to OPC in the s/s waste was 0.4. Curing of s/s waste samples took place at 20 °C and 99% humidity for 28 days. The solidified specimens were then crushed and sieved, and particles ranging in size between 425 and 850 µm were retained for the leaching tests. This size range was selected so that entire particle sections could be seen at once on the SEM at the lowest magnification. Constant pH leaching tests were conducted by mixing 30 g of s/s waste with 600 mL of ultrapure water at 25° ( 0.5 °C and by continuously controlling the addition of 5 N HNO3 to the mixture to maintain a constant pH with a variation of less than (0.1 units. The leaching vessel headspace was filled with nitrogen gas that was continuously bubbled in the leaching solution to prevent carbonation. Four different pH values (pH ) 4, 5, 6, and 7) were tested, and the duration of leaching was typically 150 h. Leachate samples were collected at various times during the tests using disposable plastic syringes and filtered through 0.22-µm membranes. After filtration, the samples were acidified with 12 M hydrochloric acid and analyzed by Inductively Coupled Plasma-Atomic Emission Spectroscopy (ICP-AES). The microstructure and composition of s/s waste particles at different stages of the leaching process were determined by SEM and EDX. More details on the experimental setup and procedures can be found elsewhere (7).
Results and Discussion Reaction of the Acid Neutralization Capacity. The extent of ANC reaction in the s/s waste was measured by monitoring
the acid consumption as a function of time (Figure 1). The reacted ANC was practically the same for tests at pH 6 and pH 7, but it was slightly higher at pH 5 and significantly higher at pH 4. The acid consumption was still increasing by the end of all the leaching tests, thus showing that the ANC was still reacting at a small rate. Figure 2 shows that the Ca release to the leachate was proportional to the acid consumption until the onset of gypsum precipitation. The proportionality constant was equal to 2.2 mol of HNO3 consumption per mole of released Ca and was independent of leachate pH. Note that the theoretical value of the proportionality constant for the reaction of Ca(OH)2 with H+ ions is exactly 2. The fact that the experimental value was slightly higher than 2 reflects the reactivity of other alkaline minerals (e.g., NaOH and KOH) in addition to calcium-based minerals. We showed earlier that the ANC of the s/s waste consists of several minerals that react sequentially (7). These reactions result in two consecutive leaching fronts advancing from the surface of the particles toward the center. Both fronts can be seen on backscattered electron (BSE) images of sections of s/s waste particles taken at two different stages of the leaching test at pH 4 (Figure 3a,b). These two leaching stages correspond to acid consumptions of 3.2 and 11.2 equiv/kg of s/s waste, respectively. Additional BSE images showing waste particles at the end of the leaching tests for all leachate pH values can be found in ref 7. The first front separates the unreacted particle core from the first shell, called shell 1, and is visible at early leaching times (Figure 3a). This front is associated with the dissolution of portlandite and the partial reaction of the CSH gel. The second front appears at a more advanced stage of leaching (Figure 3b) and is associated with the dissolution of Ca-Al hydroxy sulfates (7). The zone behind the second front is denoted shell 2. The total amount of VOL. 38, NO. 5, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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By comparison, the ANC consumption at the second front, β2, was much lower and ranged from 4476 to 7154 mol-equiv/ m3 of s/s waste. Modifications to the Shrinking Core Model To Account for the Presence of Two Separate Leaching Fronts. The basic principle of the shrinking core leach model (5, 6) is that the dissolution of the cement matrix and the release of metals occur at a single front or interface separating the unreacted core of the s/s waste specimen from an alkalinity-depleted outer shell. In addition, the advancement of the front is assumed to be controlled by the inward diffusion of acid species through the outer shell and the reaction of ANC at the interface. The specimen conversion for spherical particles, ξ, is defined as
ξ)
rp3 - rc3
(1)
rp3
where rp (m) is the initial radius of the particle and rc (m) is the unreacted core radius. The value of ξ is a function of the elapsed time t (s) after leaching starts and can be obtained by solving the following equation (5)
1 + 2(1 + ξ) - 3(1 - ξ)2/3 )
6De,HCHt rp2β
(2)
where De,H (m2/s) is the effective diffusion coefficient of H+ in the leached shell, β (mol-equiv/m3) is the ANC of the s/s waste, and CH (mol/m3) is the concentration of H+ in the leachant. The cumulative amount of ANC reaction per unit area of particle surface, MANC (mol-equiv/m2), is given by
MANC ) FIGURE 3. Backscattered electron images of s/s waste particles at (a, top) an early stage of leaching for an acid consumption of 3.2 equiv/kg and (b, bottom) an intermediate stage of leaching for an acid consumption of 11.2 equiv/kg. The leachate was maintained at pH 4.
TABLE 1. ANC Reacted at the First and Second Leaching Fronts leachate pH
ANC reacted at the first front, β1 (mol-equiv/m3)
ANC reacted at the second front, β2 (mol-equiv/m3)
4 5 6 7
23360 22131 21903 20991
6612 7154 5487 4476
reacted ANC shown in Figure 1 can therefore be decomposed into the sum of two separate terms corresponding to the ANC reaction at each one of the two leaching fronts. In the following, we explain how this decomposition was done. The ANC reaction associated with the advancement of the first front was calculated as the difference between the ANC of the core, which consists of unleached s/s waste, and the ANC of shell 1. Similarly, the ANC reaction at the second leaching front was taken as the difference between the ANC of shell 1 and that of shell 2. The ANC of each zone was evaluated from EDX measurements of Ca concentrations in the solid using the linear relationship that exists between these two parameters. The detail of these calculations is provided in Appendices 1 and 2, Supporting Information. A summary of the results is shown for each leachate pH in Table 1. The ANC consumption at the first front, β1, ranged from 20 991 to 23 360 mol-equiv/m3 of s/s waste. 1524
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Vpβξ rpβξ ) Ap 3
(3)
where Vp (m3) and Ap (m2) are the volume and outer area of the particle, respectively. When applied to a metal having initial concentration C0 (mol/m3) in the s/s waste, the cumulative amount of metal released per unit area of particle surface, MM (mol/m2), is given by
MM )
rpC0fmξ 3
(4)
where fm is the fraction of contaminant (dimensionless) that is released with the consumed ANC. Eqs 1-4 are the basic equations of the shrinking core model for spherical particles. In the following section, we modify the above equations to account for ANC reaction at two successive leaching fronts. The specimen conversion due to the advancement of the first leaching front, which separates the unreacted core from shell 1, is defined as follows
ξ1 )
rp3 - rf13
(5)
rp3
where rf1(m) is the radial coordinate of the first front. The ANC reaction during the advancement of the first front can be modeled by replacing β, the total ANC of the waste, with β1, the ANC reacted at the first front, in eqs 2 and 3. Thus,
1 + 2(1 + ξ1) - 3(1 - ξ1)2/3 ) and
6De,HCHt rp2β1
(6)
MANC )
rpβ1ξ1 3
(7)
Eqs 5-7 constitute a modified shrinking core model describing the advancement of the first front. Mechanisms Controlling the Advancement of the First Front. Figures 4a-d illustrate the results of fitting the early part of the ANC reaction data with predictions from the modified shrinking core model given by eqs 5-7. The waste particle sizes were assumed to follow a logarithmic distribution within the experimental range (425-850 µm). Since the effective diffusion coefficient of H+ in shell 1, De,H, was the only unknown in the model equations, it was used as an empirical parameter to optimize the fit. Predicted ANC reaction curves rise toward an asymptotic value that is lower than the total reacted ANC because they only account for the ANC reaction at the first front. The values of De,H that provided the best fit at each leachate pH are shown in Table 2. The calculated ratios De,H/DH, where DH is the diffusion coefficient of H+ in aqueous solution (9.311 × 10-9 m2/s) (21), are also reported. Theoretically, the ratio De,H/DH is given by (22)
ξ2 )
rp3 - rf23
where rf2 (m) is the radial coordinate of the second front separating shell 1 from shell 2. If we hypothesize that the advancement of the second front is controlled by the inward diffusion rate of H+ ions through shell 2, the relationship between ξ2 and time is given by
(8)
where φ and τ are the matrix porosity and tortuosity, respectively. Since φ < 1 and τ > 1, the ratio De,H/DH is expected to be smaller than unity. By contrast, the values of De,H/DH obtained by fitting the experimental ANC data with the modified shrinking core model range from 3.7 to 2041 (Table 2). This indicates that the ANC reacted at a much higher rate than the diffusion of H+ ions through the leached shell can account for. The large apparent increase in De,H/DH with pH is also unexpected. Therefore, contrary to one of the model assumptions, the advancement of the first front could not have been limited by the inward diffusion of H+ ions through shell 1 for the range of leachate pH conditions used in the tests. Since H+ diffusion through shell 1 is not fast enough to neutralize the OH- ions produced by the dissolution of Ca(OH)2 at the first front, then these OH- ions must diffuse outward to maintain undersaturated conditions with respect to Ca(OH)2 at the front and thus allow the front to keep advancing. Hence, it is reasonable to hypothesize that it is the outward diffusion of OH- ions through shell 1 that controls the advancement of the first leaching front. In the following, we further modify the shrinking core leach model to reflect this hypothesis. Equation 6 becomes
6De,OHCOHt rp2β1
(9)
where COH (mol/m3) is the concentration of OH- ions in pore water at the first front. Geochemical equilibrium calculations with MINTEQA2 (18) show that the OH- concentration in solution in equilibrium with portlandite at 25 °C is 19.3 mmol/L at a pH of 12.15. Using eqs 5, 7, and 9 to model the early part of the ANC reaction data with De,OH as a fitting parameter provides the same best-fit curves as those shown in Figures 4a-d. Table 3 presents the values of De,OH that provided the best fit at each leachate pH and the corresponding De,OH/DOH ratios. The diffusion coefficient of OH- in aqueous solution, DOH, was taken equal to 5.273 × 10-9 m2/s (21). The calculated De,OH/DOH are all smaller than unity and fall within the relatively narrow range of 0.016-0.034 for all leachate pH values. These results corroborate our hypothesis that the
(10)
rp3
1 + 2(1 + ξ2) - 3(1 - ξ2)2/3 )
De,H φ ) 2 DH τ
1 + 2(1 + ξ1) - 3(1 - ξ1)2/3 )
diffusion of OH- ions through shell 1 controls the advancement of the first leaching front. Mechanisms Controlling the Advancement of the Second Leaching Front and the Release of Heavy Metals. SEM/ EDX analyses of leached s/s waste particles have shown that heavy metals are only leached from shell 2 (7). Hence, the leaching of heavy metal is dependent on the advancement of the second leaching front. In the following, we extend our leach model to describe the advancement of the second leaching front and the release of heavy metals. The specimen conversion due to the advancement of the second leaching front is defined as follows
6De,HCH(t - δ)
(11)
rp2β2
where β2 (mol-equiv/m3) is the ANC reacted at the second front and the parameter δ (s) can be viewed as a time delay for the advancement of the second front. This time delay was introduced to take into account the fact that the advancement of the second leaching front is initially opposed by the neutralization reaction between OH- ions diffusing outward from the first leaching front and H+ ions diffusing inward from the surface of the particle. As the first front moves deeper inside the waste particle, the outward diffusion rate of OH- ions decreases and is eventually surpassed by the inward diffusion rate of H+ ions, thus enabling the advancement of the second leaching front. The cumulative amount of metal release per unit area of particle surface is given by
MM )
rpC0fmξ2 3
(12)
Hinsenveld and Bishop (5) define fm as the fraction of contaminant (dimensionless) that is released with the reacted ANC. We have shown earlier (7) that because of remineralization effects, only a fraction of the metals that are leached from shell 2 are actually released to the leachate. The rest of the leached metals diffuse inward and reprecipitate as metal hydroxides in a remineralization zone visible on BSE images as a thin bright layer between shell 1 and shell 2 (Figure 3b). To take into consideration the effects of remineralization on metal release, we expand fm as follows
fm )
fl fr
(13)
where fr, the remineralization factor, is defined for each metal as the ratio of the total amount of metal leached from shell 2 to the amount released to the leachate (7). The value of fr depends on the leachate pH and ranges from unity (in the absence of metal remineralization) to infinity (if all the metals that leached from shell 2 reprecipitate in the remineralization zone and none reports to the leachate). Numerical values of fr for Cu, Zn, and Pb at leachate pH 4 and 5 have been reported in ref 7. The factor fl represents the fraction of the metal concentration in the solid that is ultimately leachable, i.e. soluble in pore water at a given pH. The method of calculation of fl is presented in Appendix 3, Supporting Information. VOL. 38, NO. 5, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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FIGURE 4. Measured and predicted ANC reactions for early leaching times at (a, top left) pH 4, (b, bottom left) pH 5, (c, top right) pH 6, and (d, bottom right) pH 7. Eqs 10-13 constitute a model that describes the advancement of the second leaching front and the leaching of heavy metals. Since experimentally determined values of fr and fl are presently available only at leachate pH 4 and 5, we have limited the application of the model to these two pH conditions. Table 4 reports the experimental values of fl and fr for Cu, Zn, and Pb. The values of fl range from 0.662 to 0.706 at pH 4 and from 0.444 to 0.625 at pH 5. The higher values of fl at pH 4 are consistent with the increased leachability of metals at lower pH. Furthermore, fl is consistently largest for Zn and lowest for Cu at both pH values. The effect of remineralization on metal release is considerably larger at pH 5 than at pH 4, and fr values are consistently highest for Pb and lowest for Zn.
TABLE 2. Values of De,H and De,H/DH Providing the Best Fit for Experimental ANC Reaction Data at Early Leaching Times leachate pH
CH (mol/m3)
De,H (m2/s)
De,H/DH
4 5 6 7
0.10 0.01 0.001 0.0001
3.4 × 10-8 1.8 × 10-7 1.5 × 10-6 1.9 × 10-5
3.7 19 161 2041
TABLE 3. Values of De,OH and De,OH/DOH Providing the Best Fit for Experimental ANC Reaction Data at Early Leaching Times leachate pH
COH (mol/m3)
De,OH (m2/s)
De,OH/DOH
4 5 6 7
19.3 19.3 19.3 19.3
1.8 × 10-10 8.5 × 10-11 9.1 × 10-11 9.7 × 10-11
0.034 0.016 0.017 0.018
Measurements of cumulative amounts of Cu, Pb, and Zn released in the leachate versus time are shown in Figures 5a-c and 6a-c for leachate pH values equal to 4 and 5, respectively. Metal releases measured 10 min after the start of leaching were elevated and likely originated from the outer surface of s/s waste particles in contact with the leachant.
TABLE 4. Model Parameters for Prediction of Metal Releases δ)0 leachate pH
CH
(mol/m3)
De,H
(m2/s)
δ ) constanta
De,H/DH
De,H
(m2/s)
δ ) variableb
De,H/DH
De,H (m2/s)
De,H/DH
4
0.10
6.72 × 10-10
0.072
6.90 × 10-10
0.074
6.90 × 10-10
0.074
5
0.01
2.60 × 10-09
0.279
2.70 × 10-09
0.290
2.70 × 10-09
0.290
metal
fl
fr c
fm
Cu Zn Pb Cu Zn Pb
0.662 0.706 0.692 0.444 0.625 0.522
4.10 3.21 6.17 102 73.9 193
0.161 0.225 0.112 0.0043 0.0085 0.0027
a δ ) 3 h at pH 4 and δ ) 10 h at pH 5 for all metals. b At pH 4, δ ) 3 h for Cu, δ ) 5 h for Zn, and δ ) 6 h for Pb. At pH 5, δ ) 2.5 h for Cu, δ ) 15 h for Zn, and δ ) 20 h for Pb. c Numerical values of fr were taken from ref 7.
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FIGURE 5. Measured and predicted releases of (a, top) Cu, (b, middle) Zn, and (c, bottom) Pb at pH 4.
FIGURE 6. Measured and predicted releases of (a, top) Cu, (b, middle) Zn, and (c, bottom) Pb at pH 5.
Subsequently, metal concentrations in the leachate showed a decreasing trend for a period of time ranging from 1 to 10 h, suggesting that metal species diffused back from the leachate into the pore water of the s/s waste. This phenomenon was more apparent at pH 5, but it also occurred at pH 4. After reaching minimum values, metal releases increased again until the end of the tests, which indicates that the diffusion of metals changed direction and became oriented from the waste particle toward the leachate. Predicted metal releases are also shown in Figures 5a-c and 6a-c, and the corresponding values of the model parameters are presented in Table 4. The effective diffusion coefficient of H+ ions in shell 2, De,H, and the time delay, δ, were adjusted to optimize the fit between the predicted and experimental metal release data. The parameter De,H influences the rate at which the predicted metal release increases with time, whereas δ controls the time at which the metal starts to be released. The model predicts that metal releases always increase with time and therefore cannot predict the
early leaching behavior. Hence, the model was only used to fit the increasing trend after the minimum metal release value. Assuming δ ) 0 generally led to poor fitting of the experimental metal releases just after the minimum, although the fit improved at later times. The rationale for introducing δ in eq 11 leads us to expect that δ should decrease with decreasing leachate pH but should be constant for all metals. Taking a constant value of δ for all metals generally resulted in better agreements between predicted and measured data in comparison to δ ) 0. However, the best fits were obtained by considering that the value of δ is not only a function of pH, but also that it varies for different metals. Further research is needed to explain this result. Note that changing the value of δ only had a minor influence on the value of De,H, which was determined by fitting the later part of the metal release data. The calculated values of De,H/DH were 0.072 or 0.074 at pH 4 and 0.279 or 0.290 at pH 5 (Table 4). These values are both significantly lower than unity, as expected from eq 8, and are therefore consistent with our hypothesis that the VOL. 38, NO. 5, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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advancement of the second front is controlled by the inward diffusion rate of H+ ions through shell 2. The lower value of De,H/DH at pH 4 may be explained by the larger density and thickness of the remineralization zone at this pH (7), thus causing a greater resistance to H+ diffusion. Moreover, comparison of Tables 3 and 4 shows that the calculated values of De,H/DH in shell 2 are larger than the values of De,OH/DOH in shell 1. This implies that the matrix porosity was larger in shell 2 than in shell 1 and/or that the tortuosity was higher in shell 1 than in shell 2. These implications are consistent with the fact that shell 2 is more extensively leached than shell 1.
Acknowledgments This work was funded by the Natural Sciences and Engineering Research Council of Canada (NSERC). The authors wish to thank A. Hammond, A. Mackenzie, and A. Raitsakas for their assistance with the preparation of thin sections for SEM and with ICP analyses.
Supporting Information Available
(6) (7) (8) (9) (10) (11) (12) (13) (14) (15) (16) (17) (18)
Three appendices. This material is available free of charge via the Internet at http://pubs.acs.org.
Literature Cited (1) EPA. Solidification/Stabilization Use at Superfund Sites; EPA/ 542-R-00-010; Office of Solid Waste and Emergency Response, Technology Innovation Office: Washington, DC, 2000. (2) EPA. Solidification/Stabilization Resource Guide; EPA/542-B99-002; Office of Solid Waste and Emergency Response, Technology Innovation Office: Washington, DC, 1999. (3) Gougar, M. L. D.; Scheetz, B. E.; Roy, D. M. Waste Manage. 1996, 16(4), 295-303. (4) Batchelor, B. Environ. Eng. Sci. 1997, 14, 3-13. (5) Hinsenveld, M.; Bishop, P. L. In Stabilization and Solidification of Hazardous, Radioactive, and Mixed Wastes; ASTM STP 1240,
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Received for review July 29, 2003. Revised manuscript received December 12, 2003. Accepted December 19, 2003. ES0348400
