solidified soils under modified semi-dynamic leaching conditions

Engineering Geology 85 (2006) 67 – 74 www.elsevier.com/locate/enggeo

An evaluation of lead leachability from stabilized/solidified soils under modified semi-dynamic leaching conditions Deok Hyun Moon ⁎, Dimitris Dermatas W. M. Keck Geoenvironmental Laboratory, Center for Environmental Systems, Stevens Institute of Technology, Hoboken, NJ 07030, USA Accepted 15 September 2005 Available online 30 March 2006

Abstract Semi-dynamic leaching tests were conducted for artificial soils contaminated with lead oxide (PbO) in order to assess the longterm leaching behavior of lead (Pb). In order to simulate “worst case” leaching conditions, the semi-dynamic leaching test was modified using 0.014 N acetic acid (pH = 3.25) instead of distilled water. Lead contaminated artificial soils were prepared by mixing kaolinite or montmorillonite with fine quartz sand. The contaminated soil was then subjected to stabilization/solidification (S/S) treatment using quicklime, fly ash, or quicklime–fly ash combination. Fly ash was added in order to provide sources of aluminum, calcium, or silicate for the formation of precipitates and pozzolanic reaction products. The effectiveness of S/S treatment was evaluated by determining diffusion coefficients (De) and leachability indices (LX). A model developed by de Groot and van der Sloot [de Groot, G.J., van der Sloot, H.A., 1992. Determination of leaching characteristics of waste materials leading to environmental product certification. Stabilization and Solidification of Hazardous, Radioactive, and Mixed Wastes, 2, STP 1123, T.M. Gilliam and C.C. Wiles, eds., ASTM, Philadelphia, 149–170.] was used in order to elucidate the controlling leaching mechanisms. Slurry tests were also performed by mixing PbO with quicklime and fly ash, in order to study the immobilization mechanisms of Pb. The resulting reaction products were identified using X-ray diffraction (XRD) analyses. Overall, the test results indicated that S/S treatment was effective in immobilizing Pb and the treated soils can be considered acceptable for “controlled utilization” based on LX values. The controlling leaching mechanism was found to be diffusion, in most quicklime-treated samples. Precipitation was identified as the most likely Pb immobilization mechanism in quicklime–fly ash-treated slurries. Lead silicate, Pb2SiO4 (which is highly insoluble) is the most probable precipitate that can be associated with the decrease in Pb leachability. No evidence of pozzolanic reaction products such as calcium silicate hydrates (CSH) and calcium aluminate hydrates (CAH) was identified by XRD. © 2006 Elsevier B.V. All rights reserved. Keywords: Semi-dynamic leaching test; Stabilization/solidification (S/S); Quicklime; Lead (Pb); Leachability index (LX); Precipitation

1. Introduction Lead (Pb) has been identified as one of the most toxic elements to human health and is a widespread contam⁎ Corresponding author. Tel.: +1 201 216 8097; fax: +1 201 216 5352. E-mail address: [email protected] (D.H. Moon). 0013-7952/$ - see front matter © 2006 Elsevier B.V. All rights reserved. doi:10.1016/j.enggeo.2005.09.028

inant in many hazardous wastes sites (Lin et al., 1996). It has been reported that Pb can cause damage to the brain, red blood cells, blood vessels, kidneys and the nervous system (Lin et al., 1996; Long and Zhang, 1998). Human exposure to Pb has intensified due to industrialization. The transportation industry, which contributes leaded gasoline and Pb storage batteries, the paint industry with leaded paints, the defense industry using Pb in various

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ammunitions, the plumbing and electronics industries with leaded solder and the food industry with Pbcontaminated cans are all adding Pb to the environment. Due to the widespread use of Pb over a period of many years, the Pb loading rate in soil is approximately 20 times (or more) its natural removal rate (Nriagu, 1990). Therefore, the contamination risk to groundwater due to leaching of Pb from landfill areas and industrially contaminated land has received increasing attention (Gee et al., 1997). In this study, a stabilization/solidification (S/S) process was used in order to remediate Pb contamination, because S/S techniques have been widely applied in order to treat soils with heavy metal contamination (Conner, 1990; Yukselen and Alpaslan, 2001). During S/ S treatment, the hazardous waste potential of waste materials can be minimized by converting the contaminant into forms which are less-soluble, less mobile or less toxic and encapsulating the waste within a monolithic solid of high structural integrity (Conner, 1990). Various combinations of stabilizing agents have been used by numerous researchers in the treatment of soils contaminated with Pb. Jing et al. (2004) studied Pb leachability in cement, lime and fly ash stabilized/ solidified soil samples. Dermatas and Meng (2003) used quicklime and fly ash in order to evaluate the degree of Pb immobilization using the Toxicity Characteristic Leaching Procedure (TCLP). Li et al. (2001) studied Pb immobilization using Portland cement (OPC) and pulverized fly ash (PFA). Long and Zhang (1998) used cement in combination with various additives such as lime, fly ash, clay, apatite and silicate for treating Pb contaminated soils. Wang and Vipulanandan (1996) used Type 1 Portland cement and class C fly ash as a stabilizing agent to evaluate Pb leachability. In this study, quicklime was used as the main stabilizing agent rather than cement and hydrated lime because: a) of its economic advantage over cement and hydrated lime; b) its heat of hydration results in an accelerated rate of reaction; and c) there is limited information available to date regarding quicklime-based S/S. Upon quicklime treatment of Pb-contaminated soils, there are three possible Pb immobilization mechanisms: precipitation, inclusion and sorption. Fly ash was used as stabilizing agent alone or with quicklime in order to evaluate its effectiveness on Pb immobilization. Since fly ash contains silicon dioxide (SiO2), aluminum oxide (Al2O3) and calcium oxide (CaO), it provides calcium, aluminum, or silicate sources for the formation of precipitates and pozzolanic reaction products at high pH conditions.

Leaching is known to be a complex phenomenon because many factors may influence the release of specific constituents from a waste over a period of time (van der Sloot et al., 1996). These factors include major element chemistry, pH, redox potential, complexation, liquid-to-solid ratio, contact time, etc. (van der Sloot et al., 1996). Moreover, since very little is known about the chemical species present in waste forms and their behavior with respect to time, the long-term performance of S/S waste forms has been difficult to predict. In order to predict the long-term leaching behavior, a diffusion model is frequently used to evaluate the leaching kinetics. In this study, a diffusion theoryoriented test, the American Nuclear Society's semidynamic leaching test (ANS, 1986), was utilized in order to examine the mechanisms governing Pb leachability in quicklime-based solids. The objectives of this study are: 1) to assess the effect of clay surface area and mineralogy on Pb leachability; 2) to evaluate the effectiveness of quicklime treatment; 3) to evaluate the importance of fly ash addition in increasing Pb immobilization; 4) to determine the controlling leaching mechanisms (diffusion vs. dissolution) of Pb in treated soils; and 5) to investigate the immobilization mechanisms of Pb by using X-ray diffraction (XRD) analyses. 2. Review of diffusion model 2.1. ANS model, diffusion coefficient (De) and leachability index (LX) The long-term leachability of Pb from quicklimetreated soils was evaluated using the ANS 16.1 model (ANS, 1986). This model was established based on Fick's diffusion theory and standardized by ANS (ANS, 1986). This model can be used to determine the cumulative fraction of Pb leached against time. It has been widely reported that the leaching of the contaminant from cement-based waste forms is mostly a diffusion-controlled process (Godbee and Joy, 1974; Dutré and Vandecasteele, 1996). Due to the slow diffusion rate of contaminants, it can be assumed that a quicklime-based waste form is a semi-infinite medium, much like the cement-based waste forms examined in previous studies (Godbee and Joy, 1974; Côté et al., 1987; Andrés et al., 1995). This implies that the release of the contaminant from the waste form is negligible when compared to the contaminant's total mass. As a result of this implication, diffusion is expected to be the controlling leaching mechanism in soils treated with quicklime.

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In order to assess the long-term leaching behavior of Pb in quicklime-treated soils, the effective diffusivity (De) of the leached samples was determined based on the following ANS 16.1 model: 2
 32 De ¼ p ⦁4

 2 5 ⦁ V ⦁ Tn S ðDtÞn an A0

ð1Þ

where an = activity of a metal ion released from the specimen during the leaching interval n, A0 = total activity of a given metal ion in the specimen at the beginning of the first leaching interval, (Δt)n = tn − tn−1, V = volume of specimen in cm3, S = geometric surface area of the specimen as calculated from measured dimensions in cm2 and Tn is the elapsed time to the middle of the leaching period n in s. De is the effective diffusion coefficients (cm2/s). De values from Eq. (1) are termed “effective” because diffusion occurs in the liquid filling the interstitial spaces of a porous body. The effectiveness of quicklime, fly ash and quicklime–fly ash-based S/S was assessed by determining the leachability index (LX). According to Environment Canada (1991), LX can be used as a performance criterion for the utilization and disposal of S/S waste. When LX values are higher than 9, a treatment process can be considered effective and S/S wastes could be used in “controlled utilization”. This information indicates that the S/S wastes are acceptable for specific utilization such as quarry rehabilitation, lagoon closure, road-based material, etc. When LX values are higher than 8, S/S wastes can be used in segregated or sanitary landfills. S/S waste with an LX value lower than 8 is not considered appropriate for disposal. The leachability index is defined using the following formula: LX ¼

m 1 X ⦁ ½−logðDe Þn m n¼1

ð2Þ

where n is leaching period and m is the number of leaching periods (Godbee and Joy, 1974). 2.2. Determination of controlling leaching mechanisms In order to determine the controlling leaching mechanisms of Pb release, a model developed by de Groot and van der Sloot (1992) was used. The controlling leaching mechanism was determined based on the slope of the plot of the logarithm of cumulative fraction release, log (Bt), versus the logarithm of time, log(t) (de Groot and van der Sloot, 1992). If diffusion is

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the dominant leaching mechanism, theory suggests the following relationship: ffiffiffiffiffiffiffiffiffiffiffiffi s " ffi# 1 De ð3Þ logðBt Þ ¼ ⦁ logðt Þ þ log Umax ⦁ d ⦁ 2 p where De = the effective diffusion coefficient in cm2/s for component x (Pb in this study), Bt = the cumulative maximum release of the component in mg/m2, t = the contact time in s, Umax = the maximum leachable quantity in mg/kg, d = the bulk density of the product in kg/m3. According to de Groot and van der Sloot (1992), if the slope from Eq. (3) is equal to 0.5, Pb release will be slow and diffusion will be the controlling mechanism. At a slope of approximately 1, the controlling leaching mechanism of Pb release is dissolution. In this case, the dissolution of material from the surface proceeds faster than diffusion through the pore space of the soil matrix. Occasionally, a soluble layer exists on the surface of the material. Initially, most of the soluble material on the surface will dissolve. This process, known as surface wash-off, typically results in a slope close to 0. In most cases, following surface wash-off, contaminant release is diffusion controlled. Both dissolution and surface wash-off processes will result in the release of highly soluble materials. However, the material is not expected to be depleted during the dissolution process. 3. Experimental methodology 3.1. Preparation of artificial soils In this study, clays and fine quartz sand were combined in order to prepare artificial soils. Mixtures of clay and sand were used rather than pure clay to provide samples with a gradation more comparable to those of natural soils. Moreover, these mixtures are easier to compact than pure clay. Two types of clay minerals, kaolinite and montmorillonite, were selected in order to represent the two extremes of physicochemical clay behavior based on their surface area and cation exchange capacity (CEC). Therefore, the effects of a relatively nonreactive clay (kaolinite) on Pb immobilization were compared to those of a highly reactive clay (montmorillonite). Additionally, the amount of clay present was varied in order to evaluate its relative contribution to Pb immobilization. During the leaching test, twelve types of specimens were tested; K15L0, K15L10, K30L0, K30L10, K5C25L0, K5C25L10, M15L0, M15L10, M30L0, M30L10, M5C25L0 and M5C25L10. Letters in the specimen designation show mineralogy, i.e., K: kaolinite, M: montmorillonite, C: class C fly ash, and L:

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quicklime. Numbers following the letters indicate the percent weight of the given attributes. Since the same type of fine quartz sand was added in all mixes, sand was not included in the specimen designation. Sand content is always complimentary to the clay or fly ash content on a 100% weight basis. For example, the specimen designation K15L10 stands for 15% kaolinite and 85% fine quartz sand at a treatment level of 10% quicklime added on a weight basis (clay–sand). Similarly, K5C25L10 stands for 5% kaolinite, 25% class C fly ash and 70% sand at a treatment level of 10% quicklime. An analytical grade lead oxide (PbO) was used as the source of Pb2+ contamination. This was added to the clay–sand mixes at a concentration level of 7000 mg lead (Pb2+) per kg of solid untreated soils. 3.2. ANS 16.1 tests In this experiment, the ANS 16.1 method was modified by using a 0.014 N acetic acid solution (pH = 3.25) instead of distilled water. This modification was performed in order to simulate the possible “worst case” leaching conditions of S/S waste being disposed of in a landfill environment. In this study, specimens with a 4.0 ± 0.4 cm height and a 4.70 ± 0.05 cm diameter were prepared by compaction (in accordance with ASTM D1557-91) at optimum water content (ASTM, 1993). Following compaction, all specimens were cured for 28 days in sealed sample bags at 20 °C. Prior to the initiation of the specimen leaching test, the solids were immersed in distilled water for 30 s to rinse out any loose particles from the specimen surface. A nylon mesh harness was used in order to suspend each specimen near the centroid of the acetic acid solution in a polyethylene container. As specified by the ANS 16.1 method, the ratio of leachant volume (VL) to the specimen external surface area (S) was retained at 10 ± 0.2 cm. This ratio is maintained in order to minimize leachant composition changes as well as provide an ample concentration of extracted species for analysis (ANS, 1986). The leachate was collected and entirely replaced at designated time intervals (2, 7, 24, 48, 72, 96, 120, 456, 1128, and 2160 h). The sampled leachate was separated using a 0.4-μm pore-size membrane filter. The concentrations of soluble Pb were analyzed with a Zeeman Furnace Atomic Absorption Spectrometer (AAS) (Varian SpectraAA-400). 3.3. Slurry tests Two PbO–fly ash–quicklime slurries were prepared in order to elucidate the Pb immobilization mechanisms.

Fig. 1. Cumulative fraction of Pb leached as a function of time for untreated and fly ash-treated samples.

Fly ash–quicklime was used rather than just fly ash or quicklime to prepare the slurries since the results from the ANS 16.1 tests showed that treatment was most effective in the presence of both fly ash and quicklime. The ratio of fly ash to quicklime was 2.5 : 1 (the same ratio used previously in artificial soils). In order to evaluate both high and low levels of Pb contamination, the concentrations of PbO used were 0.25% and 8% by weight of fly ash and quicklime. A liquid to solid ratio of 20 : 1 was used. All slurries were tumbled for 24 h and then aged for 7 days. During the aging period, the slurries were shaken periodically. The leachate was filtered through a 0.4-μm pore-size membrane filter to separate solids from the leachate solution. The residue retained on the filter was air-dried and characterized by XRD analyses. The filtrate was analyzed for soluble Pb concentration. This filtration process was performed at 24 h and then at 7 days. 4. Results and discussion 4.1. Cumulative release of Pb before and after S/S treatment The cumulative Pb leachability values from untreated, fly ash-treated, quicklime-treated and quicklime–fly ash-treated specimens are plotted as a function of leaching time in Figs. 1 and 2.

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Fig. 2. Cumulative fraction of Pb leached as a function of time for quicklime and quicklime–fly ash-treated samples.

The ultimate cumulative fractions of Pb leached from all samples upon test completion are summarized and presented in Table 1. In untreated samples, regardless of composition, an increased amount of clay (from 15% to 30%) led to decreased amounts of Pb leached (Fig. 1 and Table 1). More specifically, in kaolinite–sand mixtures, although the K15L0 sample disintegrated after 7 h of testing, it was clear that a very high cumulative Pb fraction (71%) was leached out within this time period as compared to the K30L0 sample (24%) (Fig. 1 and Table 1). Similarly, in montmorillonite–sand mixtures a significant (about 50%) reduction in Pb leachability was observed when clay content was increased from 15% to 30% (Fig. 1 and Table 1). In untreated samples, montmorillonite was significantly more effective in reducing Pb leachability than kaolinite. This is probably due to the larger surface area and the greater cation exchange capacity (CEC) of montmorillonite. Upon sole addition of fly ash (K5C25L0 and M5C25L0), Pb leachability was even further decreased (less than 6%) (Fig. 1 and Table 1). This may be due to the formation of pozzolanic reaction products such as calcium silicate hydrates (CSH) and calcium aluminate hydrates (CAH). Ricou-Hoeffer et al. (2001) have shown that the formation of CSH, as a result of fly ash treatment, plays a role in the removal and stabilization of metallic cations. Therefore, it is possible that these pozzolanic

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reaction products will contribute to Pb immobilization by sorption and/or chemical inclusion. Upon quicklime treatment, Pb leachability was significantly reduced in all samples. Although the K15L0 sample disintegrated after 7 h of testing, results from the K15L10 sample, within the same time period, showed significant Pb reduction (more than 60%) (Figs. 1 and 2 and Table 1). Moreover, Pb leachability was reduced by approximately 90% in the K30L10 sample, as compared to the K30L0 sample (Figs. 1 and 2 and Table 1). Lead leachability was also significantly reduced, by more than 60%, in the M15L10 sample versus the M15L0 sample. A reduction of approximately 36% in Pb leachability was observed in the M30L10 sample versus the M30L0 sample. Samples containing montmorillonite (M15L10 and M30L10) showed very low Pb leachability (less than 1.2%) as compared to samples containing kaolinite (K15L10 and K30L10) (Figs. 1 and 2 and Table 1). Lead leachability was further decreased in quicklime–fly ash-treated samples (K5C25L10 and M5C25L10). This indicated that Pb release can be more effectively reduced by using both quicklime and fly ash rather than quicklime alone. Among all treated samples, Pb immobilization was most effective in the M5C25L10 sample (0.39% Pb release) and least effective in the K15L10 sample (35% Pb release) (Fig. 2 and Table 1). 4.2. The controlling leaching mechanisms and effectiveness of quicklime, fly ash and quicklime–fly ash treatment The controlling leaching mechanisms were evaluated by using a diffusion model (Eq. (3)) developed by de Groot and van der Sloot (1992). All slope and R2 values obtained from the diffusion model are presented in Table 2. For untreated samples (K15L0, K30L0, M15L0, and M30L0), slope values ranged from 0.05 to 0.31 (Table 2). This shows that surface wash-off appeared to be the main Table 1 Cumulative fraction of Pb leached (%) following test completion Sample

Cumulative fraction of Pb leached (%)

Sample

Cumulative fraction of Pb leached (%)

K15L0 K30L0 K5C25L0 M15L0 M30L0 M5C25L0

70.78 98.85 4.3 64.11 36.9 5.44

K15L10 35.3 K30L10 9.99 K5C25L10 1.00 M15L10 1.15 M30L10 0.87 M5C25L10 0.39

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Table 2 Regression analyses results for Pb release Sample

Slope

R2

K15L0 K30L0 K5C25L0 M15L0 M30L0 M5C25L0 K15L10 K30L10 K5C25L10 M15L10 M30L10 M5C25L10

0.05 0.31 0.43 0.21 0.19 0.4 0.51 0.6 0.64 0.71 0.63 0.6

– 0.68 0.88 0.77 0.77 0.86 0.99 0.98 0.86 0.92 0.82 0.95

controlling leaching mechanism. However, the slope values increased to approximately 0.4 for fly ash-treated samples (K5C25L0 and M5C25L0 samples) (Table 2). This indicates that diffusion was the controlling mechanism for Pb release in the fly ash-treated samples. Upon quicklime treatment, slope values ranged between 0.51 and 0.64, except for the M15L10 sample where the slope was slightly higher at 0.71 (Table 2). These slope values suggest that Pb release was controlled by diffusion. Overall, the results obtained indicate that Pb release from quicklime- and/or fly ash-treated samples was mainly controlled by diffusion. Numerous researchers have previously reported this conclusion. Côté et al. (1987) confirmed that diffusion was the main controlling leaching mechanism for Pb release in fly ash–limetreated waste. Andrés et al. (1995) have also demonstrated that diffusion was the dominant mechanism governing Pb release in stabilized steel foundry dusts. Diffusion coefficients (De) and LX values (defined by Eqs. (1) and (2), respectively) were computed and are listed in Table 3. The diffusion coefficients (De) from treated samples were significantly lower than those from untreated samples. Upon quicklime treatment, mean De values ranged from 1.02 × 10− 8 (cm2/s) to 8.98× 10− 13 (cm2/s) (Table 3). Quicklime–fly ash-treated samples (K5C25L10 and M5C25L10) had De values of 4.15 × 10− 11 (cm2/s) and 4.84 × 10− 12 (cm2/s), respectively (Table 3). According to Nathwani and Phillips (1980), diffusion coefficients generally range from 10− 5 cm2/s (very mobile) to 10− 15 cm2/ s (immobile). Therefore, it can be concluded that Pb mobility was significantly reduced upon treatment. More specifically, the K15L10 and K30L10 samples showed a De decrease of two and three orders of magnitude compared to the K15L0 and K30L0 samples, respectively. A similar decrease was observed in the M15L10 and M30L10 samples compared to their respective untreated

samples. The amount of clay also resulted in a reduction of De values for kaolinite– and montmorillonite–sand mixtures. The observed reduction was more pronounced upon quicklime treatment. Specifically, 30% kaolinite and montmorillonite samples showed De values two orders of magnitude smaller than their 15% counterparts. In samples only treated with fly ash (K5C25L0 and M5C25L0), De values were also lowered by more than two orders of magnitude, compared to untreated samples (K15L0, K30L0, M15L0, and M30L0). This decrease was even more pronounced upon quicklime addition (K5C25L10 and M5C25L10), especially for M5C25L10 sample, where a decrease of three orders of magnitude in De was noted. The type of clay was found to be an important factor that affected Pb leachability. A more pronounced reduction in Pb leachability was observed in montmorillonite samples than kaolinite samples. This is most likely due to the larger surface area and the greater CEC of montmorillonite. Since LX values for most treated samples were higher than 9, most samples can be considered acceptable for “controlled utilization”. The only exception was the K15L10 sample with an LX value of approximately 8.1. In addition, sample K30L10 had an LX value of 9.57 which indicated an acceptance of controlled utilization even though the cumulative fraction of Pb leached was approximately 10%. However, the percent Pb leached seems to have stabilized within the set period prescribed by the test protocol of 90 days. Therefore, the treatment of Pb contaminated soils with quicklime and/or fly ash mixtures was effective in immobilizing Pb. 4.3. Pb immobilization mechanisms In the low contamination slurry (0.25% PbO), four major minerals were identified by XRD: portlandite – Ca Table 3 Mean effective diffusion coefficient (De) and leachability indices (LX) Sample

Mean De (cm2/s)

Mean LX

K15L0 K30L0 K5C25L0 M15L0 M30L0 M5C25L0 K15L10 K30L10 K5C25L10 M15L10 M30L10 M5C25L10

6.12e − 06 6.07e − 07 6.84e − 10 2.24e − 07 1.02e − 07 1.59e − 09 1.02e − 08 3.62e − 10 4.15e − 11 5.55e − 11 8.98e − 13 4.84e − 12

6.55 7.73 9.74 7.76 8.07 10.04 8.07 9.57 11.02 10.33 12.64 11.98

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(OH)2, quartz – SiO2, dolomite – CaMg(CO3)2, periclase – MgO and calcium aluminum oxide sulfate hydrate – Ca4Al2SO10 · 16H2O. The main difference identified after 7 days of aging was the formation of Ca4Al2SO10 · 16H2O. The peak intensity of this mineral significantly increased after aging. Soluble Pb concentrations after 1 day and 7 days were 3.5 ppm and 2.9 ppm, respectively. In the high contamination slurry (8% PbO), six major minerals were identified by XRD: portlandite, quartz, dolomite, periclase and glaucophane – Na2Mg3Al2Si8O22(OH)2. The formation of lead silicate (Pb2SiO4) was identified, but could not be confirmed using its two strongest peaks at d-space values 2.87 and 3.13. Confirmation of this formation was not possible because it shares these two peaks with portlandite and dolomite. The formation of litharge (PbO), the Pb source, was identified after 1 day of mixing. However, it was not observed after 7 days of aging. This was probably due to the limited solubility of PbO which results in a slow release of Pb. After aging, the intensity of dolomite and glaucophane peaks increased. Soluble Pb concentrations after 1 day and 7 days were 945 ppm and 360 ppm, respectively. Overall, no evidence of pozzolanic reaction products such as CSH and CAH was observed by XRD in fly ash–quicklime slurries. Therefore, both the soluble Pb concentration and XRD results from the high contamination slurry indicate that lead silicate may be the reaction product most closely associated with a decrease in soluble Pb concentration. Palomo and Palacios (2003) found a different type of lead silicate (Pb3SiO5) when they studied fly ash contaminated with 3.125% Pb. Moulin et al. (1999) also suggested that Pb can be retained through Si–O–Pb bonds.

3.

4.

5.

6. 7.

8.

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the larger surface area and the greater CEC of montmorillonite. The sole addition of fly ash resulted in a significant decrease in the amount of Pb leached (reduced De and increased LX), compared to both untreated kaolinite and montmorillonite samples. Quicklime treatment was successful in significantly reducing the mobility of Pb (reduced De and increased LX), thus allowing only trace levels to be released. Upon quicklime and/or fly ash treatment, the controlling leaching mechanism of Pb appeared to be diffusion. In untreated samples, however, the controlling leaching mechanism was found to be wash-off. Most treated samples were acceptable for “controlled utilization” based on their LX values. No pozzolanic reaction products were identified using XRD analyses. Therefore, Pb immobilization did not seem to correlate strongly with inclusion/ sorption through pozzolanic reaction products. The controlling Pb immobilization appeared to be precipitation. The formation of Pb2SiO4 (a very insoluble compound) was observed, although not confirmed, by XRD analyses.

Acknowledgements This work was made possible by a grant (Contract No.: DE-AC21-92MC29117) from the U.S. Department of Energy (USDOE). The authors thank Dr. Mike S. Dadachov for fruitful comments related to XRD interpretation. We would also like to thank Mohammed Sharaf and Dr. Mahmoud Wazne for their critical reading of the manuscript.

5. Conclusions

References

Pb leachability in quicklime-based S/S of artificial soil samples was evaluated by performing semidynamic leaching tests. The specific conclusions pertaining to the results presented herein can be drawn as follows:

Andrés, A., Ortíz, I., Viguri, J.R., Irabien, A., 1995. Long-term behavior of toxic metals in stabilized steel foundry dusts. J. Hazard. Mater. 40, 31–42. ANS, 1986. American National Standard for the Measurement of the Leachability of Solidified Low-Level Radioactive Wastes by a Short-Term Tests Procedure. ANSI/ANS-16.1. American National Standards Institute, New York, NY. ASTM, 1993. The test method for laboratory compaction characteristics of soil using modified effort. D1557-91, Section 4, vol. 4.08, pp. 227–234. Conner, J.R., 1990. Chemical Fixation and Solidification of Hazardous Wastes. Van Nostrand Reinhold, New York, NY. Côté, P.L., Constable, T.W., Moreira, A., 1987. An evaluation of cementbased waste forms using the results of approximately two years of dynamic leaching. Nucl. Chem. Waste Manag. 17, 129–168. de Groot, G.J., van der Sloot, H.A., 1992. Determination of leaching characteristics of waste materials leading to environmental product

1. In all untreated samples, the amount of clay appears to be an important factor affecting Pb leachability. This indicated that sorption to the clay was most probably the prevailing mechanism for Pb immobilization. 2. The type of clay also had an important effect on Pb leachability. Montmorillonite samples showed a pronounced reduction in Pb leachability compared to kaolinite samples. This is most probably due to

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