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Dep. of Plant and Soil Sciences, Institute of Agriculture, The Univ. of Tennessee, P.O. Box 1071, Knoxville, TN 37901-1071
Corresponding author (messington{at}utk.edu)
Received for publication October 20, 1999.
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ABSTRACT |
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 TOPNOTES ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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Abbreviations: AMS, acid mine spoil • DDI, distilled and deionized water • EC, electrical conductivity • FA, fly ash • LSB, lime-stabilized biosolid • NFA, neutral fly ash
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INTRODUCTION |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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Studies reporting the use of FA as a possible amendment to enhance reclamation of AMS are dominated by the use of alkaline FA generated from the combustion of coal, principally from the western USA (Stehouwer et al., 1995a,b; Carlson and Adriano, 1993; Haering and Daniels, 1991; Singh et al., 1982). Using coal that has a high sulfur content and a low Ca to Fe ratio (abundant in the Appalachian region of the USA) in power generation results in the production of slightly acidic to neutral FA. Although containing little intrinsic neutralization value, these materials may be beneficial in reclamation efforts, providing essential macro- and micronutrients (K, B, Mo, and Zn) and improving AMS physical characteristics (Sutton and Dick, 1987; Carlson and Adriano, 1993). However, the additional input of a liming agent would be required for long-term acid neutralization. Stewart et al. (1997) used column leaching studies to examine the effect of eastern FA applications on coal refuse leachates. Because of the low alkalinity of the eastern FA materials, the application of FA at rates necessary to neutralize potential acidity (1.4:1 (w/w) ash to refuse mixture) was not practical. Lesser application rates (1:4 and 1:2 (w/w) ash to refuse mixtures) were effective in neutralizing the acidity of the refuse. However, near the completion of the leaching study, leachate pH began to decrease in the 1:4 (w/w) mixture. This decrease in leachate pH was also associated with a concomitant increases in soluble metals, indicating the need for additional neutralization potential.
Use of biosolids (e.g., municipal sewage sludge) has been shown to aid in the sustainable revegetation and reclamation of mined lands when applied with or without a suitable liming agent (Joost et al., 1987; Pichtel et al., 1994; Barnhisel and Hower, 1997; Cravotta, 1998). Sewage sludge is a source of organic matter, a pool of slow-release essential nutrients (N and P), and microorganisms. Despite the potential benefits associated with biosolid use in AMS reclamation, elevated concentrations of nitrate and metals have been observed in biosolid-amended AMS pore water and in ground water downgradient from biosolid-amended AMS (Cravotta, 1998). Conversely, Seaker (1991) and Sopper (1993) reported that biosolid amendments did not significantly affect downgradient ground water nitrate and metal content. If the pathogen reduction process in sewage treatment involves the addition of hydrated lime, resulting in a pH 12 material, a built-in source of alkalinity is realized. While little information is available on the use of LSB in AMS reclamation, the successful revegetation and reclamation of an AMS using LSB has been documented (Walker, 1996).
The disposal or use of NFA coupled with the utilization of LSB could result in the high volume use of NFA in land reclamation. To analyze the feasibility of LSB and NFA co-utilization for mine spoil reclamation, a field study was initiated at an abandoned strip mine site in the Cumberland Mountains near Caryville, TN. Application of LSB at rates sufficient to neutralize the potential acidity of the mine spoil was considered the reclamation mechanism, whereas NFA was applied to test spoil and vegetation response to large application rates. This field study, initiated in 1994, will require many years to evaluate the long-term chemical effect of NFA and LSB on the AMS environment during reclamation and revegetation. However, the potential long-term effect of NFA and LSB amendments can be evaluated by accelerating the weathering process through the use of simulated laboratory weathering techniques (Sullivan et al., 1986; Sullivan and Yelton, 1988; Essington, 1991; Stewart et al., 1997). The use of a simulated laboratory weathering technique, coupled with a selective sequential dissolution technique, can provide valuable insight into the fate and behavior of elements in the amended environment. The objectives of this study are to examine the potential long-term effect of co-applications of LSB and NFA to an AMS on the leaching and solid-phase speciation of minor and trace elements.
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MATERIALS AND METHODS |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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Acid–base accounting yields the net acid or neutralization potential of a material by taking the difference between the neutralization potential and the maximum acid producing potential (Sobek et al., 1978). Neutralization potential was determined in triplicate by reacting standardized 0.1 M HCl with a known mass of AMS, LSB, or NFA and boiling the solution for about 1 min. The mixture was cooled and back-titrated with standardized 0.1 M NaOH to determine the amount of HCl consumed. Acid-producing potential was determined by total sulfur analysis with a LECO (St. Joseph, MI) CR-12 furnace after sulfate removal (Grube et al., 1993). Sulfate was removed from the AMS, LSB, and NFA by filtering approximately 50 mL of an 8 M HCl solution through the material followed by thorough rinsing with DDI water (USEPA, 1974, 1978). The results of the acid–base accounting of the AMS and LSB were -12.5 and 151 kg CaCO3 Mg-1 air-dry material. Thus, the LSB amendment rate, computed on a dry-weight basis, was 88 kg LSB Mg-1 AMS (197 Mg ha-1). Further, the NFA had an acid–base account of 16.9 kg CaCO3 Mg-1 of air-dry material. Application of NFA alone at a rate of approximately 740 kg Mg-1 (1660 Mg ha-1) would be required to balance the potential acidity of the AMS.
Laboratory Weathering: Leachate Chemistry
A modified humidity cell technique was used for non-equilibrium laboratory weathering in this study (Essington, 1991). The weathering technique offers a relatively short-term mechanism for evaluating the trends in mineralogical alterations and leachate chemistry that may occur in a field environment. However, as discussed by Sullivan and Yelton (1988) and Evangelou (1995), simulated weathering only considers the relative weathering of the various components in a spoil environment. Since the degree of weathering in the field will be a function of many environmental factors (e.g., precipitation, water infiltration, oxygen diffusion, bacterial effectiveness), it is difficult to directly relate time (leaching cycle) in the laboratory to time in the field, without field validation.
A 500-g sample of AMS was placed in each of fifteen 64- x 38- x 5-cm polypropylene containers. The LSB and NFA amendment rates, as determined by acid–base accounting, were 0:0, 197:0, 197:197, 197:295, and 197:394 Mg ha-1 on a dry-weight basis. Each treatment was replicated three times. Since the objective was to use large quantities of NFA, NFA amendments were 1x, 1.5x, and 2x the calculated amendment rate for the LSB, ignoring the slight acid neutralization potential of the NFA. Although NFA rates were not based on chemical characteristics, preliminary greenhouse studies indicated that application rates of up to 224 Mg ha-1 did not adversely affect tall fescue (Festuca arundinacea Schreb.) or perennial rye (Lolium perenne L.) production.
The LSB and NFA rates were added to the 500-g spoil sample and thoroughly mixed. A 500-mL volume of DDI water was then added and the amended spoil was allowed to equilibrate for 2 h. The spoil solution was extracted from the solids by vacuum filtration through Whatman #2 filter paper into a polypropylene Erlenmeyer flask. The leachate was stored in acid-washed polypropylene bottles. The leachates were not allowed to contact glass to maintain the integrity of the leachates for boron analysis. Both the pH and the EC of the leachates were determined immediately after filtration. The remainder of the leachate was stored at 4°C for chemical analysis.
The solid cake collected on the filter paper was returned to the corresponding plastic container and distributed evenly to facilitate drying. The solids were allowed to air-dry for 1 wk, after which time the material was extracted again with a mass of water equal to the mass of the amended spoil material. This weathering technique continued for 18 cycles. Nitrate, SO4, F, and Cl concentrations in the leachates were determined by ion chromatography (IC) using a Dionex (Sunnyvale, CA) DX-100 ion chromatograph. Leachate Al, As, B, Ca, Cu, Fe, K, P, Pb, Mg, Mn, Mo, Ni, Se, and Zn concentrations were determined using a Model 61 Thermo-Jarrell Ash (Franklin, MA) inductively coupled argon plasma–optical emission spectrometer (ICP–OES).
Sequential-Selective Dissolution and Elemental Analysis
Sequential-selective dissolution analysis was performed on samples from the 2- and 18-wk weathering cycles to determine the effect of the LSB–NFA amendments and weathering on the mineral distribution of various trace elements in the amended AMS. The procedure used in this study was developed by Stover et al. (1976) and revised by Lund et al. (1980) for the indirect characterization of trace element solid-phase speciation in biosolid (sewage sludge) and biosolid-amended soil. This method was selected over those that are applicable to strictly mineral systems and are purported to have better reagent selectivity for specific mineral pools (Tessier et al., 1979; Ahnstrom and Parker, 1999). Since the specificity of reagents for specific mineral pools in complex multicomponent systems is tenuous at best, we employed the procedure that was developed for biosolid-amended systems. Further, the results are interpreted to indicate relative lability and the transformation of metals from soluble to sparingly soluble phases during weathering, rather than to substantiate the existence of specific metal pools.
The sequential extraction procedure of Lund et al. (1980) partitions trace elements into the following operationally defined pools: exchangeable-soluble (F1), adsorbed (F2), organic (F3), carbonate (F4), and sulfide (F5) forms. In addition, trace elements residing in the residual (non-extractable) fraction were computed by the difference between the total metal content and the sum of the extractable metals. Extractions were performed in 50-mL polypropylene centrifuge tubes using triplicate 2-g samples with 25 g of the appropriate reagent. The metal pools of the solid phase were extracted by use of the following reagents sequentially: F1, 0.5 M KNO3 for 16 h; F2, DDI water for 2 h repeated three times and the extracts composited; F3, 0.5 M NaOH for 16 h; F4, 0.05 M EDTA for 6 h; and F5, 4 M HNO3 for 16 h. The samples were centrifuged (1500 x g) between steps for approximately 10 min to separate the solid and liquid phases. The extracts were analyzed by ICP–OES for Co, Cr, Cu, Ni, Pb, and Zn. The mass of a given element extracted by a given reagent was calculated as follows: µg extracted = 
- 
where A is the concentration in µg g-1 (liquid) extracted, B is the concentration in µg g-1 (liquid) extracted from the previous step, and C is the mass in grams of extracting solution left over from proceeding extraction step and carried to the current step (Sposito et al., 1982).
The total elemental content of the AMS, LSB, NFA, and the Cycle 2– and Cycle 18–amended AMS was determined in triplicate using the microwave-induced fluoroboric acid dissolution technique of Nadkarni (1984) with modification (Ammons et al., 1995). A 200-mg sample (<60 mesh) was pretreated with 2 mL HF in polyallomer centrifuge tubes, mixed by vortexing, and allowed to react for 16 h. After pretreatment, 5 mL of aqua regia (3:1:1, HCl to HNO3 to H2O) was added to each tube and vortexed. Samples were then heated in an 800-W microwave for a 3 min reaction time at 80% power to hasten the dissolution reaction. After cooling, 1 g of boric acid was added to each sample to neutralize excess HF present. Samples were vortexed and heated in the microwave for an additional 10 min at 20% power to dissolve the boric acid. While still warm, the samples were again vortexed to facilitate boric acid dissolution. The cooled samples were filtered through Whatman #42 filter paper into 100-mL volumetric flasks and brought to volume with DDI water. The chemical analysis was performed using ICP–OES for Al, As, Ba, Ca, Cd, Co, Cr, Cu, Fe, K, Mg, Mn, Mo, Na, Ni, P, Pb, S, Se, Si, and Zn. The total elemental contents of the NFA, LSB, and AMS are unremarkable, as concentrations fall within the ranges or near the mean values established by Mattigod et al. (1990) and Eary et al. (1990) for NFA, Essington and Mattigod (1990) for anaerobically digested sewage sludge, and Bowen (1979) (reports mean elemental content of shales) and Adriano (1986) (reports ranges for shales and clays) for AMS (Table 1).
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RESULTS AND DISCUSSION |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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The concentrations of NO3 in the LSB-amended AMS leachates were similar to those observed in the unamended AMS leachates. Nitrate concentrations in biosolid-amended AMS pore and ground waters are generally higher than NO3 concentrations in unamended AMS waters, even though the mean NO3 concentration in amended AMS is similar to that in unamended AMS leachates (Cravotta, 1998). The pulse of soluble NO3 that occurred around Cycle 15 in both the amended and unamended AMS systems mimics the observed reduction in leachate pH (Fig. 1). Maximum proton and nitrate production occurs during Cycle 14 for the 0 and 197 Mg ha-1 FA-amended AMS, and during Cycle 15 for the 295 and 394 Mg ha-1 FA-amended systems. Further, higher NO3 concentrations and greater proton activity are observed in the 295 and 394 Mg ha-1 FA-amended systems. Clearly, FA influences the nitrification process, although it is not readily apparent as to the mechanism. Apparently, by Cycle 11 the population of nitrifying organisms was sufficient to oxidize organic N in both the unamended AMS (residual coal) and LSB-amended AMS (residual coal and LSB-born organic N), releasing NO3 and protons. It is not clear, however, as to why nitrification occurs to the same extent and during the same weathering cycles in both the unamended and amended AMS, particularly when one considers the drastically differing pH conditions (pH 3–3.5 versus pH 7–7.5) and the character of the available organic N (coal versus biosolid). The oxidation of pyrite or other FeII–bearing minerals, coupled with NO3 reduction, may be responsible for the decreasing NO3 concentrations as weathering proceeds beyond Cycle 15. However, the trends in soluble Fe and SO4 established prior to Cycle 11 were not influenced, and proton activity returned to pre–Cycle 11 levels.
Increasing B concentrations with increasing NFA amendment rates appeared to be the only apparent effect of NFA on leachate chemistry. Biosolid amendments resulted in higher B concentrations during the initial weathering cycles in the AMS leachates, relative to the unamended AMS leachates. However, NFA amendments influenced leachate B concentrations throughout the weathering study, and unlike the leachate concentrations of the other elements, B concentrations remained relatively stable (B regeneration). Average (and range) leachate B concentrations in the NFA-amended AMS during Cycles 3 through 18 were 0.008 mmol L-1 (0.007–0.009 mmol L-1), 0.011 mmol L-1 (0.009–0.013 mmol L-1), and 0.013 mmol L-1 (0.011–0.015 mmol L-1) for the 197, 296, and 394 Mg ha-1 treatments. According the Keren and Bingham (1985), the B threshold concentration for irrigation water is 0.027 mmol L-1 for very sensitive crops. While it is apparent that the NFA regenerates solution B, concentrations are still below levels that would adversely affect even the most sensitive plants. Other elements that are of potential concern in FA and FA-amended systems, such as As, Mo, and Se, were below detectable levels during the study (<0.1 mg L-1 for As and Se and <0.01 mg L-1 for Mo).
Selective Sequential Dissolution and Elemental Analysis
The concentrations of trace elements in NFA, and Cu and Zn in LSB exceed their respective metal concentrations in the AMS (Table 1). Biosolid application did not significantly affect the Co, Cr, and Cu content of the AMS (Table 2). The Ni and Zn content of the LSB-amended AMS was greater than the unamended material, whereas the Pb content decreased. Fly ash applications increased the total metals content of the LSB-amended AMS, particularly at the highest application rate. The observed metal concentrations in the LSB- and NFA-amended AMS, relative to the unamended AMS, are generally consistent with the metal contents of the individual materials: high NFA or LSB metal content, relative to AMS, enhanced metal levels in the amended AMS, whereas low NFA or LSB metal content, relative to AMS, diluted metal levels in the amended AMS.
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As with all selective dissolution techniques, the results obtained provide only an indication of the mineralogical pools in which trace elements reside. For the objectives of this study, a knowledge of the distribution of trace elements in these pools is sufficient. Any change in trace element distribution that results from NFA and LSB amendment will be evident. Further, and perhaps more significantly, the sequential extraction results will provide an indication of the relative stability (lability) of the metal-bearing phases, as affected by weathering and the various treatments.
Fly ash amendment of AMS and weathering had relatively little effect on the solid-phase speciation of Cr and Pb (Table 3). Both elements were found in relatively stable phases. In general, greater than 90% of the total Cr was found in the residual fraction, with approximately 5% found in the F5 fraction, irrespective of treatment or extent of weathering. A significant, but minor (1 to 3%) decrease in the percentage of Cr in the residual fraction is evident in all systems as a result of weathering. In the unamended AMS, F4 Cr increased with weathering. The Cr content of the remaining fractions was not significantly affected by treatment or weathering. Lead was primarily found in the F4 and residual fractions, averaging 20 and 74% of the total, respectively. Approximately 4% of the total Pb was found in the F5 fraction. The distribution of Pb in these three fractions was not influenced by weathering, and treatment effects were minor. The stability of the Cr- and Pb-bearing phases illustrated by the selective extraction results is consistent with results of the leachate analyses: both Cr and Pb were below detectable levels throughout the study.
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In the unamended and Cycle 2 AMS, approximately 70% of the total Cu as found in the F1 fraction. At the completion of the weathering study, the F1 fraction contained 21% of the total Cu and the residual fraction contained 33%. Copper was the only trace element found to have a significant presence in the F3 pool, presumably associated with the organic matter. The percentage of Cu in the F3 fraction was not affected by weathering and relatively small increases in the F4 and F5 fractions occurred. Copper in the LSB-amended AMS was found principally in the F3 and F4 fractions. Similar to the unamended AMS, Cu was not present in the residual fraction. Upon weathering of the LSB-amended AMS, residual Cu increased to 55% of the total while the Cu content of the F1, F3, and F4 fractions decreased. With the inclusion of NFA as an amendment, there was an apparent, but insignificant increase in the residual Cu pool with increasing application rate in the Cycle 2 materials. None of the Cu-bearing mineral pools were significantly affected by NFA application. Upon weathering of the NFA-amended materials, residual Cu increased and the percentage of Cu in the F1 and F3 fractions decreased. The distribution of Cu in the various chemical pools of the Cycle 18 NFA- and LSB-amended AMS was statistically similar, irrespective of NFA rate.
With the exception of Cu, the application of LSB and NFA to AMS had little effect on the solid-phase speciation of the trace elements in the Cycle 2 and Cycle 18 materials. Further, the speciation of Cr and Pb was not influenced by the degree of weathering. Excluding Cr and Pb, weathering of the unamended and LSB- and NFA-amended AMS resulted in the formation of less labile phases, at the expense of the labile F1 phase. In the unamended AMS, weathering resulted in a decrease in F1 Co, Cu, Ni, and Zn and an increase in residual Cu and Co and F5 Ni and Zn. In LSB-amended AMS (with or without NFA), weathering decreased the percentages of Co, Ni, and Zn in the F1 fraction with a concomitant increase in the percentages of these elements in the F4 fraction. The application of NFA appeared to incorporate a significant amount of residual Cu into the Cycle 2 material, a fraction that increased upon weathering. Despite the transformation of many of the trace elements to less labile solid phases, weathering only affects a small percentage (<10%) of the total amounts present (excluding Cu).
The application of LSB was responsible for reducing the solubility of various trace (Cu, Ni, and Zn) and major (Al and Fe) metals, as evidenced by the chemical composition of the AMS leachates (Fig. 3). The transformation of metals from relatively labile pools (F1 and F3) to less labile pools (F4 and residual) is evidence of the trace element transformation that might be predicted given the alkaline conditions of the LSB-amended AMS leachates and the high concentrations of Ca and SO4 (and potentially CO3). These conditions are similar to those found in weathered alkaline coal FA environments for which the mineral phases have been documented (Mattigod et al., 1990). Using equilibrium solubility concepts, Reddy et al. (1988), Eary et al. (1990), and Schwab (1995) have predicted the formation of sparingly soluble metal oxides and hydroxides (Al, Fe, Cu, and Ni), ferrites and silicates (Zn), sulfates (Cu), and carbonates (Cu, Ni, and Zn).
The use of sewage sludge and other biosolids, in conjunction with a suitable acid-neutralizing material, is an established mechanism for promoting the revegetation and reclamation of acid mine spoil. In this study, LSB was employed to neutralize the potential acidity of AMS and to provide the reference environment to examine the effect of high-volume NFA co-applications on AMS leachate chemistry and trace metal mineral pools. Based on the results of the laboratory weathering study, the loading of NFA onto AMS during reclamation with LSB would have little effect on leachate composition. Indeed, the increased leachate pH, salinity, concentrations of common salts, and the decreased concentrations of hydrolyzable (Al, Fe, and Mn) and trace (Cu, Ni, and Zn) elements primarily responded to the LSB application alone. Only leachate B concentrations were influenced by NFA applications; increasing with increasing NFA rate. Although below threshold levels for sensitive crops, leachate B was regenerated by the NFA, such that concentrations were stable during the 18 weathering cycles. Other potentially problematic elements common to FA (As, Mo, and Se), were below detectable levels in all NFA-amended AMS leachates. The application of LSB also resulted in the incorporation of a small but significant pool of relatively labile trace elements. However, weathering of the LSB-amended AMS resulted in the significant reduction of the labile trace element pool and the formation of relatively less labile pools. The application of NFA did not affect the partitioning of the trace element into the various mineral pools. Based on the results of the simulated weathering study, the co-application of large volumes of NFA with LSB appears to be a suitable mechanism for the disposal of this combustion by-product. Irrespective of NFA application scenario (use or disposal), B will be of potential concern.
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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