Determining controls on element concentrations in fly ash leachate

Waste Management & Research (1995) 13, 435-450

DETERMINING CONCENTRATIONS

CONTROLS IN FLY

ON ASH

ELEMENT LEACHATE

E. J. Reardon I, C. A. Czank~, * C. J. Warren I, R. Dayal 2, and H. M. Johnston 2 tDepartment of Earth Sciences, University of Waterloo. Waterloo. Ontario, N2L 3GI. Canada and 'Environmental Sciences Department. Ontario Hydro, Toronto, Ontario. M8Z5S4, Canada

(Received 2 December 1993, accepted in revisedJorm 12 August 1994) It is possible, by conducting fly ash leaching tests at two different water:ash ratios, to determine whether or not the concentration of an element in the leachate is controlled by mineral solubility. If a mineral solubility control exists, the element's concentration can be readily predicted with chemical equilibria models. On the other hand, if the concentration of all element is controlled by the rate of release From ash particles, more sophisticated hydrochemical transport models are required. The conduct of leach tests on waste materials at two different water:solid ratios is recommended as a general procedure in waste management. In this study, the concentration of Na, K, CI, B, and Cr(VI) in Lakeview fly ash leachate were found not to be controlled by mineral solubility. Therefore, the solution concentrations of these elements will be influenced by the relative amounts of water to ash in a mix, and the flux of these elements from hydrating primary ash particles with time. Evidence of solubility control was found for Ca, Sr, SO4, A1, Si, As(V), and Se ill leachate solutions. Calcium and SO4 concentrations were controlled by gypsum solubility; Sr, probably by a (Ba,Sr)SO4 solid solution; A1 and Si, by a hydrous aluminosilicate, probably allophane: Mg, by a hydrous magnesium silicate like sepiolite; and As, by either a ferric metal arsenate or strongly associated with a ferric oxide phase. The probable control on Se concentrations is through coprecipitation in secondary sulphate mineral precipitates. Key Words

Fly ash, leachates, solubility control, chemical equilibria.

I. Introduction N u m e r o u s studies have categorized the composition o f water generated as a result o f its contact with fly ash. Leachate waters can have markedly different compositions depending on the source of coal, flue gas conditioning process, design o f the combustion system, and whether lime or limestone injection processes are implemented for desulphurization. F o r example, total dissolved solid concentrations (TDS) m a y vary from hundreds to tens o f thousands o f mg 1-t; leachate pH values can vary from alkaline to acid and both major and trace element concentrations can vary over orders o f magnitude. Even a single sample o f fly ash can show marked differences in leachate water chemistry depending on reaction time and water:solid ratio in batch equilibrations, or column length and flow rate in dynamic leaching tests. The mineral and glass phases that constitute fly ash material are formed over a wide range o f temperatures in the stack environment o f a coal-fired electric generating station * Present address: Environ Corporation, Carnegie Center, Princeton, N J, U . S . A .

0734 242X/95/050435 + 16 $12.00/0

© 1995 ISWA

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E . J . Reardon et al.

(Hansen et al. 1984; Furuya et al. 1987). For the most part, they are oxidized, or partly oxidized, anhydrous phases. However, when fly ash is brought into contact with water at earth surface conditions, all of these phases are unstable, i.e. they will dissolve and more stable and less soluble secondary phases will precipitate. Despite their high solubilities, some of the primary phases of fly ash, especially the glass and crystalline aluminosilicate particles, dissolve very slowly. In addition, secondary hydrous aluminosilicate alteration products are very insoluble and build up as alteration rinds on the surfaces of the primary phases. This further impedes the dissolution of the primary phases as the mass flux of ions and water between these phases and the porewater becomes diffusion-controlled. Besides the slowly-reacting aluminosilicate phases, fly ash also contains minerals formed from the condensation of elements at lower temperatures in the stack environment. These minerals often become attached to the surfaces of rising glass particles. These include sulphate, borate, fluoride and chloride minerals that typically have high solubilities and rapid dissolution rates. Oxide phases, such as CaO, MgO, AI,O~, Fe_~O3 and Fe304 are also present. These phases show a wide range in solubility and dissolution rates. The extreme heterogeneity in particle compositions, solubilities and dissolution kinetics contributes to the great variability in observed leach test results, depending on reaction time and the water:solid ratio used in the test. The dominant chemical characteristic of a leachate, however, is usually established within minutes after addition of water. This results from a flush of ions to the solution from the readily soluble minerals present in the ash. Secondary minerals may precipitate immediately depending on the ion concentrations attained during this initial flush of ions. The aqueous phase thus acts as a medium for the transfer of elements from the set of anhydrous, metastable phases of the original ash to a set of hydrous, more stable secondary precipitates in the reacting water/ash mix. After several tens of minutes, dissolution and precipitation reactions slow substantially and a relatively stable solution composition is attained (Kopsick & Angino 1981). At this point, the constituents leached from the ash can be classified as to whether or not solid phases control their concentration levels in solution. The prediction of elemental concentrations in fly ash leachates hinges on differentiating between these two classes of dissolved constituents. Simple chemical equilibria models can be applied to predict the species concentrations of elements controlled by mineral solubility under both static and dynamic flow conditions, but more complex models that incorporate mineral dissolution kinetics and diffusion are required to predict the concentrations of elements not controlled by mineral solubility. Their concentrations will be a complex function of the characteristics of the hydrologic flow regime: water:solid ratios; diffusivities of reaction rinds and secondary coatings; and the grain size distribution of the fly ash. Determining whether the concentration of a particular element in fly ash leachate is controlled by mineral solubility or not presents problems. The quantities of secondary mineral precipitates that may form are often too low to detect by X-ray diffraction analyses. Generally, the only evidence of solubility controls is gained from calculations of mineral saturation indices with the use of chemical equilibria programs such as WATEQ4F (Ball & Nordstrom 1991). Due to the limitations and uncertainties in the thermodynamic database of these programs, this indirect approach rarely provides unequivocal evidence to establish solubility control for a particular element. In this study, the utility of fly ash leaching studies using different water:solid ratios to provide direct evidence for or against mineral solubility control on elemental concentrations is

Determining controls" on element concentration

437

TABLE 1 Chemical composition of Lakeview fly ash (after Johnston & Eagleson 1989) Element AI As B Ba Ca CI Cr Fe(lil)

pg g- *

Element

pg g-*

113,000-114,512 76.1-89 78 360-700 28,857-34,000 50 137-170 80,671

Fe (total) K Mg Na S Se Si Sr

84,147-89,300 12,784 5698 3710 2683 5.06 203,047 53

demonstrated. This information can aid considerably in the interpretation of mineral saturation index calculations from equilibria programs. In this report, the results for a series of timed water/fly ash equilibration experiments conducted at two water:solid ratios are presented. The data is used to discern whether mineral solubility controls exist for particular elemental concentrations.

2. Material The fly ash material used in the leaching experiments was obtained from the Ontario Hydro (Canada) Lakeview thermal generating station. Electrostatic precipitators are used to remove as much as 99.9% of fly ash from the flue gas stream prior to discharge of the flue gas. The sample represents dry material collected from the electrostatic precipitators at the station. The grain size of the ash ranged between 0.5-150 pm with a mean diameter of 6-9 pm. Spherical particles, some of which are hollow, predominate over the entire size range. Carbon particles in the ash tend to be coarser and more angular than other particles. The specific gravity of the fly ash varies widely from one particle to another, ranging from less than 1.0 for hollow spheres (termed cenospheres or floaters) to greater than 3.0 for magnetic spinels, with an average value of ~2.4 (Dayal et al. 1992). A complete chemical analysis of the material was reported by Johnston and Eagleson (1989) and the data relevant to this study are recorded in Table 1.

3. Experimental A series of batch leaching experiments were performed on the Lakeview fly ash. The experiment involved reacting samples of the ash in 3:1 and 6:1 water:ash ratios for up to 32 days in 60 ml high density polyethylene reaction bottles. The bottles were mounted horizontally on a carousel and immersed in a constant temperature bath maintained at 25+0.5°C. The carousel was rotated by a rotisserie motor at 2.5 rpm during the reaction period. A sufficient number of replicate samples for each water:ash mix were reacted to provide all the solution samples that were obtained for chemical analyses during the 5-week experiment. Each reaction bottle provided sufficient solution for only one complete set of analyses. Solution samples were obtained at weekly intervals and filtered

E. J. Reardon et al.

438

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Fig. 1. X-ray diffraction analysis results for reacted and non-reacted Lakeview fly ash samples. Diffractogram shown at lop of the diagram represents the intensity difference between reacted and non-reacted samples. Peak labels: G, gypsum: Q, quartz: M, mullite: S, magnetite (spinel).

through 0.2 pm membrane filters. Each sample was diluted 1:2 with nano-pure water and split into two portions. One portion was acidified with one drop of concentrated HNO3 and set aside for cation analyses on a Varian Atomic Absorption Spectrophotometer (AAS). The unacidified portion was reserved for anion analyses on a DIONEX AS3 ion chromatograph. Alkalinity was measured on the unacidified sample with standardized HCI using a Metrohm auto-titrator. Arsenic was analysed by AAS using a hydride generation technique (Siemer & Koteel 1977). The following colorimetric techniques were used to analyse for AI, Si and B using a Beckman DU-6 Spectrophotometer. Aluminum was analysed using a pyrocatechol violet method (U.K. Department of Environment 1979); Si by a molybdate blue method (Technicon 1973); and B by an azomethine-H method (John et al. 1975). Solution pH was monitored throughout the reaction on a separate replicate sample that was not used for chemical analysis to avoid K and C1 contamination by the reference electrode. Samples of the reacted and non-reacted fly ash were analysed by X-ray diffraction using Cu Kz radiation generated at 30 mA and 40 kV. Specimens were step-scanned as random powder mounts from 3-63 o 20 at 0.05 o 20 steps; integrated at l s s t e p '.

4. X-ray diffraction analyses of reacted and non-reacted fly ash material X-ray diffraction analyses provided little useful information in this study. Despite the complexities in the behaviour of the solution compositions with time, X-ray analysis yielded virtually identical results for reacted and non-reacted samples. Representative diffractograms for unreacted Lakeview fly ash and material recovered from the 3:1 water:ash run after 4 weeks of reaction are shown in Fig. 1. The fresh ash sample contained the minerals quartz, mullite, and magnetite (spinel)

439

Determining controls on element concentration 9.4 9.2

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Time (days)

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Fig. 2. Solution pH and alkalinity values for different water:fly ash ratios during 32 days of reaction. 7-1, 3: I water:fly ash results; 0 . 6 : 1 water:fly ash results.

based on the X-ray diffraction data. The X-ray diffraction pattern for the treated ash sample was similar to the pattern for the fresh ash except that the absolute intensities for all peaks had increased due to the preferential removal of glassy material covering the mineral phases (Dudas & Warren 1987) and the appearance of a peak corresponding to a d-spacing of 0.756 nm suggesting the presence of gypsum in the sample. Included in Fig. 1 is a plot of the intensities of the reacted sample minus those of the nonreacted material. The difference plot shows that gypsum (CaSO4.2H,_O) was indeed the only identifiable secondary mineral phase produced during the reaction of fly ash with water. X-ray diffraction, however, is often not sensitive to the development of small, but potentially important, quantities of secondary mineral phases, The normal cut-off limit for the detection of crystalline phases in powdered mounts by X-ray diffraction analysis is on the order of 5% by mass. 5. Elemental concentrations and other solution properties vs. time

5.1 p H and alkalinity The evolution o f p H and alkalinity values with time for the 3:1 and 6:1 water:ash mixes are shown in Fig. 2. There are two opposing processes which establish the pH value for a fly ash sample (Furr et al. 1977). The dissolution and hydrolysis of oxide components, such as CaO and MgO, contributes to an increase in solution pH: CaO + H , O ~ C a '-+ + 2 O H -

[1]

Offsetting the pH increase, contributed by the oxide components, is the dissolution of soluble acids, such as B203; and salts containing hydrolyzable constituents, such as Fe2(SO4)3 and A12(SO4)3. The relative quantities of these soluble bases and acids in a particular ash establish whether the leachate will have a dominant acid or basic character. In most cases, the quantity of soluble oxides outweighs the quantity of soluble acid phases and fly ash material generates an alkaline pH within minutes of reaction with water. In contrast to the rapid dissolution of soluble acids and bases, which establishes the overall initial acid/basic character of the leachate, there is the subsequent slow dissolution of silica-rich glass from ash particles. This reaction is promoted at both high and low

E. J. Reardon et al.

440 0.4

0.3 I

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Time (days) Fig. 3. Concentration o f CI in the water/fly ash extracts. [Z], 3:1 water:fly ash results: O, 6:1 water:fly ash results.

pH and, in the case of alkaline leachates, results in a decrease in pH with time due to the hydrolysis of silicic acid: SiO: + 2 H 2 0 ~ H4SiO]~ H.~SiO4 + H +

[2]

For tile Lakeview ash in Fig. 2, tile dissolution of soluble acids and bases generated initial pH values greater than 9, and the slow dissolution of silica-rich glass accounts for tile gradual decline in pH during the 32-day reaction period. Although the differences are small (less than 0.1 pH units), one curious feature of the pH data is that the 6:1 water:ash mix exhibits systematically higher values during the reaction period than the 3:1 mix. This is inverse to the alkalinity results which are consistent with an expected greater alkaline character imparted to the leachate when more ash is added to the water, i.e. at lower water:ash ratios. Since H_,BO~ is the principal component of tile alkalinity, the lower pH, associated with the lower water:ash ratio, probably represents greater base neutralization when more B_,O~ is able to dissolve.

5.2 ChlorMe

The CI concentration in the water/fly ash extracts are presented in Fig. 3. The chloride results indicate that solution concentrations are nearly constant with time, and that doubling the mass of solid relative to the mass of water, doubled the Cl concentration in solution. These two observations indicate that CI in ash is in the form of readily soluble salts, probably attached to surfaces of glass particles formed at low temperatures in the stack environment (Hansen et al. 1984), and that no secondary mineral precipitates control its concentration in solution.

441

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Fig. 4. Concentration of Na. K, Cr and B in the water/fly ash extracts. []. 3:1 water:fly ash results; O, 6:1 water:fly ash results.

5.3 Sodium. potassium, boron and chromium The concentrations of Na, K, B and Cr in the extracts are shown in Fig. 4. For all four elements, 50% or more of the final concentrations (after 32 days of reaction) were attained within the first few hours of reaction, indicating that readily soluble salts are important sources of these elemental constituents. Each element displayed a behaviour similar to CI as indicated by a doubling of the dissolved concentration upon halving the water:ash ratio. This indicated the absence of secondary mineral solubility controls on these elements. The difference with respect to the CI results, however, was that after the initial flush of ions to the solution, there were small, but progressive, increases in the concentration of Na, K, and B with time. This indicated the existence of an important reservoir for these elements within the unreacted anhydrous cores of the primary fly ash particles, which was slowly released by diffusion with time (Warren & Dudas 1984). The concentrations of B released in these leaching experiments were high ( l - 4 m m o l l '; 10-40mgl-~). These levels are higher than those observed for Na, K and Mg. Only the concentrations of Ca and SO4 in the leachates exceeded B levels. Boron is also the principal component of alkalinity. Our calculations indicate that H,BO~ generally comprises more than one-half of the measured alkalinity. Although vascular plants exhibit a considerable range in tolerance for B, concentrations in excess of 0.5 mmoll ' (5.3 m g l - ' ) are usually detrimental (Keren & Bingham 1985; Warren 1992). Our observed release levels of 1~, mmol 1-' are for 6:1 and 3:1 water:ash ratios. Presumably even greater concentrations of B would be released from Lakeview fly ash

442

E. J. Reardon et al.

at lower water:solid ratios, more representative of field disposal conditions. However, this was not evident in a study by Vorauer (1989). Groundwater samples taken from two Lakeview landfill sites in that study showed that B concentrations do not exceed 2 mmoll ~ (21 mg 1-~). Detailed monitoring of a sump piezometer at one of these sites between 1982 1987, however, showed steady increases in the concentrations of B. Consequently, as originally indicated by Johnston & Eagleson (1989), B should be considered an element of concern for the disposal of Lakeview fly ash. The concentration of Cr in the leach solutions (200-600~tg 1-~) was two orders of magnitude greater than the solubility of Cr(OH)3 at these pH conditions ( ~ 5 fig I-~) (Baes & Mesmer 1976; Rai et al. 1987). This indicated that the leached Cr must, principally, be in the more soluble, hexavalent CrO]- form. The lack of evidence for solubility control on leachable Cr from fly ash was also reported by Dayal et al. (1992) in a study of material from the Nanticoke generating station in southern Ontario. The quantity of Cr leached in these experiments represents only about 1% of the total Cr in the original ash (Table 1). The dominant valence state of Cr in fly ash is expected to be Cr(llI), not Cr(VI), since it is principally associated with the ferromagnetic fraction (Hansen et al. 1984; Warren & Dudas 1988). So although the leached results in this study indicate a lack of solubility control on Cr(VI) concentrations in solution, Cr(lll) concentrations are likely controlled by the solubility of Cr(OH)3 or a mixed (Cr, Fe)(OH)~ solid solution (Rai et al. 1987).

5.4 Calcium, sttvntiton and sulphate

Concentrations of Ca, Sr, and SO4 showed similar temporal variations but very different behaviour from the previously considered elements (Fig. 5). The most notable feature was that doubling the mass of material relative to the mass of water did not result in a doubling of the solution concentrations for these elements. This indicated a solubility control on their solution concentrations. Both X-ray diffraction analyses (Fig. 1) and calculated mineral saturation indices (SI) (Fig. 6) indicated gypsum solubility control on Ca and SO4 concentrations. Strontium results for the 3:1 and 6:1 mixes also suggested solubility control. Although celestite (SrSO4) is a potential solubility-controlling phase, it is more likely that a (Ba,Sr)SO4 solid solution controls the concentration of Sr. There are two reasons for this. First, the SI calculations indicate undersaturation of the solutions with respect to celestite (Fig. 6). Secondly, the total concentration of Ba in fly ash is typically higher than Sr, which is the case with the Lakeview fly ash used in this study (Table 1). Assuming a similar source for these two elements in the original ash (likely oxides) and similar dissolution behaviour (both oxides are readily soluble), more Ba would enter the solution phase than Sr when the water is initially added to the ash. Since the solubility product for BaSO4 is three orders of magnitude lower than that for SrSO4 (10 9.97VS. 10-t"('3), virtually all of the leached Ba would precipitate as barite. This is substantially more than the total possible amount of celestite that could precipitate, owing to its higher solubility and the lower Sr concentration in the ash. Since Sr easily substitutes for Ba, substantial quantities of Sr could be removed from the solution in the form of a (Ba,Sr)SO4 solid solution. The solubility of this solid solution would be considerably lower than that of celestite (Ba concentrations were all below detection in this study) and thus would explain the calculated undersaturation of all solutions with respect to celestite.

Determining controls on element concentration

443

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Fig. 5. Concentration of Ca, Sr, and SO4 in the water/fly ash extracts. I-1, 3:1 water:fly ash results; I , 6:1 water:fly ash results.

Identical leaching behaviour in lysimeter samples from a fly ash field in Pennsylvania were reported by Fruchter et al. (1990). Their results indicated that fly ash porewaters were consistently undersaturated with respect to pure celestite and they ascribed this undersaturation to the solubility control of either a Sr-substituted barite or a Basubstituted celestite. They also conducted laboratory experiments in the mixed BaSO4SrSO4-H20 system and obtained similar findings.

5.5 Silicon, aluminum and magnesium

The concentration of Si, AI, and Mg in the water/fly ash extracts are shown in Fig. 7. The virtual independence of the concentrations of these elements with water:ash ratio is a clear indication of solubility control of these elements. For A1 and Si, the likely solid phase control is a hydrous aluminosilicate such as allophane or imogolite (Warren & Dudas 1985). These minerals are common alteration products of primary silicates and volcanic ash in weathering environments. Allophane forms through the hydrolysis of volcanic glass at high Si concentrations and pH values, while imogolite forms from the alteration of allophane at lower Si and pH conditions (Wada 1987). Although imogolite was not contained in the WATEQ4F database, the saturation indices with respect to allophane, as calculated by WATEQ4F, are shown in Fig. 6. These results indicate close correspondence to saturation for all solutions. Based on the ion activity

444

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product for allophane [IAP=(H4SiO4°)-'-(AI(OH)#)-'.(H+)2], increases in the concentrations of AI and Si with decreasing pH is consistent with an allophane solubility control. The importance of the formation of secondary silicates and their solubility control was further supported by the results of a l-week 20:1 water:ash equilibration experiment performed to gauge the potential quantity of Si released to solution. The concentration of Si attained in this experiment was the same as in the 6:1 and 3:1 experiments: ~ 4 0 p m o l l -I ( l . 2 m g l '). The results for the concentration of Mg also indicate solubility control in the 3:1

445

Determining controls on element concentration 70

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and 6:1 mixes. Hydroxide, carbonate and sulphate phases were all ruled out as solubility controls based on WATEQ4F saturation index calculations which indicated a substantial degree of undersaturation with respect to these phases. The likely control on Mg is a hydrous silicate or aluminosilicate. Sepiolite [Mg4Si60~5(OH)2"6H20] is known to form at similar temperature and pH conditions (Stoessell 1988). Figure 6 records the calculated saturation indices for sepiolite. Although slight undersaturation is indicated, it is noted that the SI calculations are based on the pure endmember. Aluminum can substitute for Si in sepiolite to produce a more stable (and less soluble) solid phase. The concentration of Mg sharply increases between the 20- and 32-day sampling for both reaction mixes. Similar results have been obtained in other studies (Hodgson et al. 1982) and is consistent with the expected solubility behaviour of sepiolite with decreasing pH. However, substantial variability may have arisen as a result of the sampling procedure. Each point on the concentration plots represents a sacrificed sample. This means that the 32-day sample was taken from a different bottle than the earlier samples and pH measurements were taken from a separate sample again. This was done to avoid continually opening a reaction mix during the reaction period so as to avoid sample contamination. Consequently, caution must be exercised in detailed interpretation of point-to-point variations. 5.6 Arsenic and selenium

The independence of the concentration profiles for As and Se with respect to water: ash ratio indicated solubility control for these elements (Fig. 8). Only the total

E.J. Reardon et al.

446

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Time (days)

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Fig. 8. C o n c e n t r a t i o n o f As and Se in the water/fly ash extracts. [-I. 3:1 water:fly ash results: O , 6:1 water: fly ash results.

concentration of these elements in the leachates was determined in this study, and not the individual oxidation states. In the case of As, As(V) is usually found to be the dominant form (Turner 1981; Silberman & Harris 1984; Hansen et al. 1984). However, even if substantial quantities of As(Ill) species are leached from fly ash, Johnston & Eagleson (1989) note that they would be oxidized to As(V) species in the presence of oxygen within several days (Cherry et al. 1979). There are many highly insoluble arsenate minerals that form in low temperature environments and potentially control arsenic concentrations in leachate solutions. It is clear from Fig. 6, however, that tricalcium arsenate [Ca.~(AsO4)2-4H20], the stable phase for these pH conditions, does not control As concentrations in the leachate solutions. Arsenic may be controlled by precipitation of one of the highly insoluble transition metal arsenates, such as ferric arsenate. This possibility cannot be evaluated because iron concentration in the leachates were below detection (less than 0.02 mg I-~). In most natural environments, As(V) species are strongly attenuated by ferric oxides. Johnston & Eagleson (1989) reviewed the available literature on this process and suggest it as a probable control on As concentrations in fly ash leachates. Differentiating between sorption and solubility control is very difficult, so sorption of As(V) by ferric oxides could perhaps be described equivalently by solubility with respect to ferric arsenate.

Two studies have examined the valence state of Se in fly ash leachates. Both van der Sloot et al. (1985) and Niss e / a l . (1993) found that Se(IV), in the form of SeO~- ions, was the dominant valence state in fly ash leachates. Selenium(Vl) was found to be dominant over Se(IV) species in only one out of four fly ashes examined by Niss et al. (1993). Our data indicate that there is no pure phase solubility control for Se in solution. All possible alkaline earth and transition metal selenite minerals are substantially more soluble (Smith & Martell 1976) than the observed Se concentrations in the leachates. Both Johnston & Eagleson (1989) and Dayal et al. (1992) report similar Se concentrations for both batch and column leach tests of Lakeview fly ash (30-180~gl-~). One likely control suggested by Johnston & Eagleson (1989) is co-precipitation with sulphate minerals, such as gypsum and barite, into which SeO4-'- can readily substitute (Ticknor et al. 1988) and perhaps SeO3-~- as well. These seem reasonable solid phase controls on the concentration of Se observed in this study.

Determining controls on element concentration

447

6. Discussion

One of the principal problems that confronts a researcher interested in the controls on elemental concentrations in leachates of waste materials is whether mineral solubility exerts control on a particular element. The information is extremely useful since chemical equilibria models can be used to predict leachate concentrations of elements controlled by mineral solubility. When no solubility control exists, the concentration of an element will be controlled by reaction kinetics and transport properties, which require more complex predictive models. The most reliable evidence to establish mineral solubility control is by X-ray diffraction analyses or direct observation of the leached product by scanning electron microscopy. Unfortunately, these techniques are rarely helpful because the solubility-controlling phases are often not highly crystalline or present in low quantities. Unless a controlling mineral phase occurs in excess of 5 wt %, for example, X-ray diffraction analysis is unlikely to detect its presence. Therefore X-ray diffraction analysis will be of little use in detecting solubility-controlling phases for trace elements. This study demonstrates that direct evidence of mineral solubility control is provided by the results of leaching tests conducted at two different water:solid ratios. Simply put, if an element's concentration does not double when the water:solid ratio is halved, there must be a solid phase control on that element's concentration in solution. Despite the usefulness of this information, it is disappointing that less than 10% of papers on fly ash leaching provide test results at more than one water:ash ratio. In this study, some simple fly ash leaching experiments conducted at 6:1 and 3:1 water:ash ratios provided evidence of solubility control on Ca, St, SO4, Se, As(V), Si, Mg, and AI concentrations; and a lack of solubility control on Na, K, CI, B, and Cr(VI) concentrations. This information can aid substantially in the interpretation of mineral saturation index information generated by chemical equilibria programs routinely used in waste management. As a further illustration of the utility of this approach, sample results from a study of cement carbonation by Reardon & Dewaele (1990) are presented in Fig. 9. This figure records the concentration variations of four elemental constituents in cement porewater during the progressive reaction with carbon dioxide. The carbonation of a cement/water slurry results in a systematic decrease in pH and a transformation of all the original minerals in the cement into new phases. The results clearly reveal A1 as an element for which mineral solubility controls existed throughout the carbonation reaction (various calcium aluminum silicates); Na, as an element for which no control was present; K, for which a solubility control developed during the reaction (Kassociated Si-gel); and SO4, for which a solubility control existed at the beginning of the reaction (ettringite), but became more soluble and dissolved completely as a result of the progressive decrease in pH. Finally, there is one caveat in applying this concept to discern solubility controls on particular elements in leaching experiments. One conclusion drawn in this study is that if an elemental constituent does not double in concentration with a halving of the water:solid ratio, then a solid phase control must exist for this element. In this study, all elemental examples in which solubility control was inferred happened to exhibit similar concentration levels for the two water:solid ratio leach tests. However, this may not always be the case. Consider, for example, as discovered in the present study, that C1 appears to have no solubility control in a leaching experiment. If some trace constituent in the leachate were controlled by the solubility of its chloride salt, e.g.

448

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AgCI, then there would be an apparent doubling in the Ag concentration with a doubling in the water:solid ratio (i.e. a halving in the leachate C1 concentration). This is imposed because of the constraint of the solubility product relation at equilibrium: K,,, = [Ag +][C1-1

[31

Fortunately, this doubling in Ag concentration could not be construed to indicate the absence of solubility control because it is inverse to the concentration change expected. However, it does illustrate that solid phase control on an element can be manifested in different behaviours of the concentration difference results between the two water/ solid leaching tests. A solid phase control on a particular elemental constituent causes deviations l¥om the concentration doubling effect associated with the absence of solid phase control. However, the extent and direction of this deviation will be a function of the complexities of the overall solution composition. 7.

Conclusions

Experiments were conducted to determine the leaching characteristics of Lakeview fly ash with time and to determine which elemental concentrations in leachate are controlled by mineral solubilities. The following conclusions were made: - - F o r 3:1 and 6:1 water:solid ratios, the solution concentration levels of Na, K, CI,

Determh~hTg controls on element concentration

449

B, and Cr(VI) are not controlled by mineral solubility. Therefore, the solution concentrations of these elements, at water:ash ratios greater than 3:1, will be influenced by the relative amounts of water to ash in a mix, the quantity of readily soluble material containing these elements, and the flux of these elements from hydrating primary ash particles with time. - - T h e concentrations of Ca, Sr, SO4, AI, Si, As(V), and Se in leachate solutions are controlled by mineral solubility. Consequently, their solution concentrations can, for the most part, be determined by solubility relations incorporated in chemical equilibria models commonly used in waste management. Calcium and SO4 concentrations are controlled by gypsum solubility; Sr, probably by a (Ba,Sr)SO4 solid solution; AI and Si, by a hydrous aluminosilicate, probably allophane; Mg, by a hydrous magnesium silicate, such as sepiolite; and As, by either a ferric metal arsenate or strongly associated with a ferric oxide phase. Although there is strong evidence for solubility control on Se concentrations, a pure selenite phase is unlikely. It is more likely that Se concentrations are controlled by co-precipitation in secondary sulphate mineral precipitates. The concentrations of B released in the leaching experiments ( l ~ m m o l l ~) are higher than those observed for Na, K and Mg. Only Ca and SO4 concentrations exceed B levels in the leachates, making B the third most abundant element in Lakeview fly ash leachate and the principal component of the alkalinity. - - T h e undertaking of leaching tests on waste materials at two different water:solid ratios is recommended as a general procedure to assist in determining solubility controls on elemental constituents, and to aid in the interpretation of the results of chemical equilibria programs.

8. Acknowledgements The authors wish to thank both Ontario Hydro and the Ontario Ministry of the Environment for their funding of this project. The excellent analytical work of Ron Kcllerman and the staff of Waterloo's Water Quality Laboratory is greatly appreciated.

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