Environ. Sci. Technol. 19g2, 26, 507-5 11
Tysklind, M.; Rappe, C. Photolytic Transformation of Polychlorinated Dioxins and Dibenzofurans. Tenth International Symposium on Chlorinated Dioxins and Related Compounds; Bayreuth, Germany, September 1990. Crosby, D. G.; Wong, A. S. Science 1977,195,1337-1338. Liberti, A,; Brocco, D.; Allegrini, I.; Cecinato, A.; Possanzina, M. Sci. Total Environ. 1978, 10, 97-104. Eitzer, B. D.; Hites, R. A. Environ. Sci. Technol. 1989,23, 1389-1401.. Finlayson-Pitts, B.; Pitts, J. Atmospheric Chemistry: Fundamentals and Experimental Techniques;John Wiley and Sons, Inc.: New York, 1986. Korfmacher, W. A.; Wehry, E. L.; Mamantov, G.; Natusch, D. F. S. Enuiron. Sci. Technol. 1980, 14, 1094-1099.
(19) Behymer, T. D.; Hites, R. A. Environ. Sci. Technol. 1988, 22, 1311-1319. (20) Dulin, D.; Drossman, H.; Mill, T. Environ. Sci. Technol. 1986,20, 72-77. (21) Orth, R. G.; Ritchie, C.;Hileman, F. Chemosphere 1990, 18, 1275-1282. (22) Leighton, P. A. Photochemistry of Air Pollution;Academic Press: New York, 1961.
Received for review April 8, 1991. Revised manuscript received August 2, 1991. Accepted October 21, 1991. This work was supported by the U. S. Department of Energy through Grant 87ER- 60530.
Leach Studies on Hazardous Waste Monoliths from Continuous Solidif ication/Stabilization Processing Mlchael R. Powellt and R. Mahalingam"
Department of Chemical Engineering, Washington State University, Pullman, Washington 99 164-27 10 The continuous solidification/stabilization processing of hazardous wastes by microencapsulation in a polyester matrix has been shown to be feasible in our laboratories, through the use of a static mixer/reactor. Leach rates for the solidified samples were observed to be more a function of emulsion droplet size and the volume fraction of waste incorporated into the emulsion than the degree of initiator mixing as quantified by the number of mixing elements placed between the initiator injection point and the exit of the mixing tube. More importantly, the leach rates were substantially lower than those reported for solidified matrices from grout processing and urea-formaldehyde binding.
Introduction Solidification/stabilization ( S / S ) processes are intended for use in situations where chemical detoxification of a hazardous waste stream is either not possible or prohibitively expensive. In solidification/stabilization disposal of hazardous wastes, the approach is to treat the waste stream in such a way as to enable it not only to be safely transported, stored, or buried, but also to reduce the mobility of its toxic constituents. S/S technologies currently employed are grout processing and urea-formaldehyde binding, and these appear to be based more on economics than on the delivery of desirable leach characteristics. The polyester encapsulation process (1)is an example of a novel solidification/stabilization technique. In this processing technique, the waste stream is mixed with a commercially available polyester-in-styrene binding material (2) until the waste is finely dispersed within the resin. For an aqueous waste, such mixing results in a stable waste-in-resin emulsion. The resin phase of this emulsion is next polymerized, resulting in the waste becoming encapsulated within the high-strength cross-linked resin shell. Background In the polyester encapsulation process, the first step involves the emulsification of the aqueous waste into the polyester resin, to a maximum possible 7:3 volume ratio. +Present address: Battelle Pacific Northwest Laboratories, Richland, WA 99352. 0013-936X/92/0926-0507$03.00/0
Typical emulsion droplet sizes generated are on the order of 2-10 pm. In the next step, small amounts of a freeradical initiator are added to initiate the polymerization of the continuous resin phase. A high-strength monolithic solid results within -10 min, the time interval being controllable by various process parameters. Batch processing parameters and the types of wastes (low-level nuclear wastes; chemical wastes-heavy metals, organic chlorides, pigment sludges, decontamination chemicals; mixed wastes; etc.) treated are reported elsewhere (1,3-7). The gross physical characteristics of the encapsulated waste monolith are reported in ref 1. The current research program demonstrates the successful conversion of the batch process into a continuous one, through the use of a static mixer/reactor (8,9);here, the two steps are carried out sequentially in different sections of the same static mixer/reactor. The leach characteristics of the solidified matrix resulting from this process are evaluated and interpreted in this paper.
Experimental Section The experimental unit for the solidification/stabilization part of the studies consisted of four basic sections. The first three were the flow arrangements for the resin, simulated waste, and polymerization initiator, and the fourth was the stainless steel static mixer/reactor. A schematic of the unit is shown in Figure 1. The static mixer was obtained from Komax Mixing Systems (Wilmington, CA) in the form of 100 individual mixing elements. A fixed number of these elements were packed by hand into the 0.3191-in.-I.D. (0.375-in.-o.d.) stainless steel tube, also supplied by Komax. The water-extensible resin used was Aropol WEP-661P polyester resin (Ashland Chemical, Columbus, OH). This resin contains -60% styrene, 3540% polyester, and proprietary amounts of surfactants and polymerization promoters (cobalt naphthenate and dimethylaniline) (see ref 1 for more on the chemistry of this resin). Although the process is amenable to a variety of wastes as stated earlier, the simulated waste solution used in the present runs was a 20 wt % sodium sulfate solution. This sodium sulfate solution is the principal constituent in low-level boiling water nuclear reactor wastes (IO). The polymerization initiator used was a solution of methyl ethyl ketone (MEK) peroxide dissolved in methyl
0 1992 American Chemical Society
Environ. Sci. Technol., Vol. 26, No. 3, 1992 507
G
L
400
,
i
C
,
L L i
F S 7 e (i-r-) Flgure 2. Sample drop size distribution from phase-contrast microscopy: $ = 0.4, QE = 135 mL/min, D, = 2.31 pm, D,, = 3.44 pm. Drop
Figure 1. Schematic of S/Sexperimental unit. Key: (A) N,-blanketed resin reservoir, (B) simulated-waste reservoir, (C) N,-pressurized MEK peroxide reservoir, (D) nitrogen cylinder, (E) solidification container, (F) Komax static mixer, (G) resin metering pump, (H) MEK peroxide rotameter, (I) simulated-waste rotameter, (J) pressure gauge, (K) check valves, (L) polymerization initiator injection port locations.
ethyl ketone (Delta X-9, Pennwalt). All solidification/ stabilization runs were carried out with all three process streams held at a temperature of 20 2 "C. Full details are available in refs 8 and 9. The solidification/stabilization runs were next followed by experimental work designed to relate the leach characteristics of the encapsulated waste to the emulsion drop size and the degree of initiator mixing into the emulsion. The basis for this approach was the results from our previous batch processing studies ( I ) . The drop size used for characterizing the emulsion was the Sauter mean emulsion droplet diameter, D32,defined as
*
where f ( D )is a drop size distribution function which gives the number fraction of drops that are of size D , and this in turn was correlated with the number of Komax mixing elements (N) needed for forming a stable emulsion, the volume fraction of the waste incorporated into the emulsion (4), and the volumetric flow rate of the emulsion through the mixer (QE). The degree of initiator mixing into the emulsion was defined by the number of mixing elements (N2)placed between the initiator injection port and the exit of the rnixer/reactor. The number of mixing elements N2was varied by the appropriate placement of the initiator injection taps. One injection tap was installed at 15 mixing elements from the mixer exit and another at 30. For each set of mixing elements N2,volume fraction of waste 4, and emulsion Sauter mean diameter D32 values, three samples were cast in the form of 3-cm-diameter right cylinders with the length-to-diameter ratio L I D = 1.0. One of the casts was used for the leach test, another was used for scanning electron microscopy (SEM) studies, and the third was kept as a spare. A Hitachi S-570 scanning electron microscope was used to determine the average cell size in each of the solidified samples, after cleaving them to expose the encapsulated waste cells. For purposes of comparison with the emulsion drop sizes, the cell sizes from SEM measurements were recalculated to be also expressed as DS2,following a procedure described in refs 8 and 11. The American Nuclear Society (ANS) 16.1 static leaching procedure for hazardous wastes was used to evaluate the degree of waste immobilization in each of the samples. This test was selected over the toxicity characteristic leaching procedure (TCLP) and the EP toxicity 508
Environ. Sci. Technol., Vol. 26, No. 3, 1992
tests because the ANS 16.1 test gives substantially more information about the rate at which waste leaches out of the solidified product. The TCLP and E P toxicity tests are used primarily in the determination of whether a waste treatment process meets the EPA requirements for land disposal of solidified wastes. The Materials Characterization Center MCC-1 test (developed by Battelle PNL) is similar to the ANS 16.1 test in that it gives quantitative leach rate information, but the MCC-1 test is most applicable to cases where the waste is expected to be in contact with stagnant groundwater. The ANS 16.1 test is preferred because it simulates the more intense leaching caused by mobile groundwater (12, 13). The ANS 16.1 static leaching procedure is conducted as follows. The sample is cast into some type of simple geometry (e.g., right cylinder, cube, sphere) and placed initially into a volume of reagent grade water (minimum resistance 0.2 MO/cm) equal to 10 f 0.2 cm times the sample surface area (cm2). Next, the sample is transferred into fresh deionized water after 1,2,4,7, and 14 days (total of 28 days). The ANS 16.1 test for hazardous wastes is a modified version of the ANS 16.1 standard test for radioactive wastes, which is conducted for a total of 90 days. One sample from each of the three solids from each run was suspended in the appropriate quantity of deionized water (resistance 21'7.0 MO/cm). A thin piece of nickel wire was used to keep each sample suspended in the leachate. The leaching of all samples was carried out in 500-mL Nalgene bottles, which are made of nonreactive polypropylene. That neither the bottles nor the nickel wire contributed to the concentration of the sodium ion in the leachate was verified by analyzing the leachate from both a bottle containing only deionized water and a bottle containing deionized water and 20 cm of nickel wire. These leachates were found to contain no detectable traces of sodium ion. Capture of sodium ion by the polypropylene was determined to be insignificant by a test in which the sodium ion concentration of a 0.000 04 M (initial concentration) sodium sulfate solution stored in a polypropylene bottle was found to be constant over a period of 3 weeks. The leachate was replaced with fresh deionized water at the recommended intervals. The removed leachate was analyzed for sodium ion by an atomic absorption spectrophotometer (AAS; Perkin-Elmer 2280). Due to the leachate dilution procedure required prior to analysis and the AAS and the small drift in the sodium lamp voltage, the leachate concentrations are estimated to be accurate to within *5%.
Results and Discussion A sample drop size distribution (DSD) histogram for the emulsion, as determined by the phase-constant microscope
Table I. S/S Processing Parameters
sample
&, mL/min
9
,V
.V2
D I L ,pm
z1
100 125 125 100 650 400 650 400 400 400
0.40 0.55 0.55 0.70 0.40 0.55 0.40 0.55 0.45 0.45
50 50 50 30 50 50 50 50 95 95
15 15 30 30 15 15 30 30 15 30
9.2 7.6 7.7 42 3.6 4.3 3.6 6.0 2.4 2.8
22 23 Z6 27 28 z9 z10 Zll 212
0
0 003. 0
Table 11. Leachability Index" and Effective Sodium Ion Diffusivities* Computed from Experimental Leach Rates
sample Deff,i Deff,s Z1 22 23 Z6 27 28 Z9 Z10 Zll
212
Deff,7
Deff,id
10
_- _-
25
20
15 T i r e (days)
4 = 0.70.
Flgure 5. Cumulative leachability vs time, for
Li
Deff,28
19.8 3.2 4.0 2.6 2.0 16.3 49.2 21.2 139.0 72.9 6.3 38.8 12.4 192.0 72.3 601.0 541.0 446.0 244.0 437.0 0.0056 0.0053 0.149 0.0267 29.8 4.0 1.9 0.94 6.2 48.3 0.0056 0.0041 0.021 10.7 0.092 1.9 6.0 1.9 161.0 14.8 0.017 0.150 0.037 18.8 0.216 0.0061 0.039 0.011 11.3 0.098
"Units of L;, none. *Units of Dpff,X
5
11.4 10.4 10.5 9.36 13.1 11.3 13.3 11.1
12.7 13.1
cm2/s.
--____I
2 OC3. 5
0
10
20
15 T i l e (days)
25
jd
Flgure 6. Cumulative leachability vs time, for q5 = 0.45, __-
10 T
l
0
0N = 3 0
I
I
ON=53 A h=72 A N=95
W
6
0 0 0 J
I G
c
I
o
-euL8--n A--A-
lp o o 0
10
I
I 1
-0-
I
1
n
' A
A
~
5
w
-3-
-31
15 Time (days)
25
20
33
Flgure 3. Cumulative leachablllty vs time, for q5 = 0.40. 0
0 L v
---
I
c 203
-
0 000
0
c 623
----I :t i c 3
Flgure 7. Minlmum attainable Sauter mean droplet diameter.
diffusivities and leachability indexes. The formulation of the cumulative leachability, the leachability index, and the effective diffusivity are described below. If an is defined as the cumulative mass of Na+ leached out during the nth extraction, then the cumulative leachability Ln of a sample is calculated from .JL
00 0 0 I -
0
5
10
15 Time (days)
23
'
L
25
I 30
Flgure 4. Cumulative leachability vs time, for q5 = 0.55.
(PCM),is presented in Figure 2; the distribution is seen to be log-normal. The various process parameters for the solidified matrices and the resulting D,,'s calculated from SEM cell size measurements are summarized in Table I. The results of the ANS 16.1 static leach tests are given in Figures 3-6 and Table 11. Figures 3-6 are plots of the cumulative leachability of each sample as a function of time, while Table I1 presents calculated effective Na+ ion
n
Ln = (C a n / A J / ( S / V ) i=l
(cm)
where A, is the total mass of Na+ originally present in the solid cylinder and S and V are respectively the surface area and volume of the cylinder. The effective diffusivity of the leached contaminant (Deff)is calculated per ANS 16.1 standard as
where T is defined as Environ. Sci. Technol., Vol. 26, No. 3, 1992
509
T = (0.5 (tn1I2- t,-llIz))z
(s)
/,c---!----o
C 602 r .
and At, is the duration of the nth leaching interval, t, tn-l, in seconds. The leachability index is defined per ANS 16.1 standard as
where P = 1.0 cmz/s. The leachability index is convenient in that it is a means of quantifying the leach characteristics of a sample by a single number. The cumulative leachability data presented in Figure 3 indicates that, at a volume fraction of waste $I = 0.4, the leach characteristics are relatively independent of 0 3 2 . On this basis, it can be concluded that very favorable leach characteristics could be expected for samples encapsulated at volume fractions of 0.4 provided that 032 is equal to or less than -12 pm. Thus, it appears that the extra energy expended to reduce D3, from 12 to 5 Km or less does not yield a significant improvement in leach characteristics, at volume fraction 4 = 0.4. Figure 4,on the other hand, shows the much stronger dependence of the leach charat volume fraction 0.55. In this case, acteristics on 032, the energy expended to reduce the drop size does result in a significant improvement in the leach characteristics. Whether the extra energy expenditure is economically justified depends on whether the larger drop sizes result in contaminant immobilization to a satisfactory degree. The cumulative leachability data for volume fraction 0.7 are presented in Figure 5. The most important feature of these data is the large increase in the magnitude of the leach rate that is observed as the volume fraction of the waste is increased from 0.55 to 0.70. Much of this increase is probably due to the resulting larger drop sizes for these samples. In cases where the solidified matrix Sauter mean cell diameter 0 3 2 and volume fraction are essentially held constant and only the number of mixing units N2 is varied, the higher N2 value gives the product with the lower leach rate (Figure 6). The data in Table I1 indicate that the principal mode of contaminant leaching is the removal of surface contamination and not diffusion. If diffusion was the primary driving force for contaminant removal, then the effective diffusivity, Deff,values should be constant with respect to time. For all samples, the effective diffusivity is seen to steadily decrease with time. From a process economics standpoint, the data presented in Figure 7 appear quite informative. Because the major operating cost of the encapsulation process is likely to be the cost of the resin (14), it is desirable to operate the process a t the highest allowable 4. As 4 is increased, however, (D32)min must also increase. But, it has been determined that as the drop size increases, the leach rate of the encapsulated waste also increases; hence, prior to selection of the 4 a t which to operate, the maximum tolerable leach rate must be specified. These studies have shown that by encapsulating waste at relatively low volume fractions (say, 4 5 0.5) and small Sauter mean sizes (D32 -< 5 pm) extremely low leach rates can be obtained. Another key feature of the polyester microencapsulation process is seen in Figure 8. It shows the leach characteristics are lower than those typically reported for grout or urea-formaldehyde (UF) processing (15). Note that the grout and UF leachability data are based on lithium ion as the tracer. Lithium is expected to have a higher ionic mobility than sodium ion, but based on leach tests using other ions (e.g., cesium), the cumulative leachability de510
Environ. Sci. Technol., Vol. 26,
No. 3, 1992
E
0
13
20 ;!re
30
CI
(days:
Figure 8. Comparison of leachability: polyester encapsulation vs grout processing vs UF binding (grout/UF data from ref 15).
termined using lithium is expected to be less than 10% greater than that determined by using sodium ion as the tracer. Conclusions The polyester process, whether operated in a batch or continuous mode, is seen to provide a high degree of waste immobilization, without sacrificing the leach characteristics. Also, the process gives significantly less leaching than either grout or UF processing. Any leaching that might occur is merely the washoff of the surface contamination. Acknowledgments Useful discussions with John J. Barich, 111, of the U.S. Environmental Protection Agency, Seattle, WA, are very much appreciated. Registry No. Na2S0,, 7757-82-6;Aropol WEP-661P, 3733877-5.
Literature Cited (1) Subramanian, R. V.; Mahalingam, R. In Toxic and Hazardous Waste Disposal; Pojasek R. B., Ed.; Ann Arbor Science Publishers: Ann Arbor, MI, 1979; Vol. I, Chapter 14. (2) Technical Data for A r o p o p WEP 661P; Ashland Chemical Co.: Columbus, OH, 1974. (3) Mahalingam, R.; Juloori, M.; Subramanian, R. V.; Wu, W.
P. Proceedings o f the National Conference o n Treatment and Disposal of Industrial Wastewaters and Residues, Houston, TX, April 26-28, 1977; p 9. (4) Biyani, R. K. M.S. Thesis, Washington State University, Pullman, WA, 1978. ( 5 ) Mahalingam, R.; Biyani, R. K.; Shah, J. T. Znd. Eng. Chem. Proc. Design Dev. 1981, 20, 85. (6) Mahalingam, R.; Jain, P. K.; Biyani, R. K.; Subramanian, R. V. J . Hazard. Mater. 1981,5, 7706. (7) Bailey, W. J.; Mahalingam, R. J . Hazard. Mater. 1981,5, 145. (8) Powell, M. R. M.S. Thesis, Washington State University, Pullman, WA, 1990. (9) Powell, M. R.; Mahalingam, R. Znd. Eng. Chem. Res., in press. (10) Krischer, W.; Simon, R. Testing Evaluation and Shallow Land Burial of Low and M e d i u m Radioactive Waste Forms; Harwood Academic Publishers: New York, 1984. (11) Powell, M. R.; Mahalingam, R. J. Dispersion Sci. Technol., in press. (12) Measurement of the Leachability of Solidified Low-Level Radioactive Wastes by a Short-term Test Procedure: 1986; ANSI/ANS-16.1; American Nuclear Society: Chicago, IL, 1986. (13) Background S t u d y on t h e Department of a Standard Leach Test; EPA-600/2-79-109; Environmental Protection Agency, U S . Government Printing Office: Washington, DC, 1979.
Environ. Sci. Technol. 1992, 26, 511-516
Mahalingam, R.; Biyani, R. K.; Jain, P. K.; Subramanian, R. V. A Techno-Economic Evaluation o f Current Hazardous Wastes Solidification Processes. Report 78/13/27; WSU College of Engineering, 1978. Filter, H. E. In Toxic and Hazardous W a s t e Disposal; Pojasek, R. B., Ed.; Ann Arbor Science Publishers: Ann
Arbor, MI, 1979; Vol. I, Chapter 11.
Received f o r review February 25, 1991. Revised manuscript received October 10, 1991. Accepted October 24,1991. M.R.P. was supported on a U N O C A L Fellowship. Material support came f r o m Komax Mixing Systems and Ashland Chemical Co.
Chemical Stability and Decomposition Rate of Iron Cyanide Complexes in Soil Solutions Johannes C. L. Meeussen,* Meindert G. Keizer, and Frans A. M. de Haan Department of Soil Science and Plant Nutrition, Wageningen Agricultural University, P.O. Box 8005, 6700 EC Wageningen, The Netherlands
In order to improve the assessment of the bioavailability and behavior of cyanide in the environment, the speciation of dissolved cyanide was studied under pH and redox conditions relevant to soil and groundwater environments. The partition of cyanide over free cyanide [HCN(aq) + CN-] and iron cyanide complexes [or hexacyanoferrates, e.g., Fe(CN)63-and Fe(CN),4-] at thermodynamic equilibrium was calculated as a function of pH and redox potential. These calculations show that the free cyanide form will predominate a t chemical equilibrium in the soil. In groundwater from sites contaminated with cyanide, however, only complexed cyanide was found, indicating that the speciation of cyanide is determined not by chemical equilibrium but by decomposition kinetics. In daylight, iron cyanide complexes appeared to decompose rapidly (ca. 8%/h). In the dark, the rate of decomposition appeared to be much slower and was proportional to the fraction of hexacyanoferrate present as HFe(CN),3- and H2Fe(CN)2-. The decomposition rates of both species were determined as a function of temperature and used to model decomposition kinetics. Good predictions were made of decomposition rates under various pH and redox conditions. Introduction Cyanide is a highly toxic chemical which commonly occurs as an industrial contaminant of soils. Although cyanide is produced in small amounts by many different organisms ( I ) , it reaches toxic levels in the environment solely because of human activities. Cyanide is used in several industrial processes because of its ability to form stable complexes with a range of metals. It was and still is used for this purpose in the mining industry, the metallurgical industry, and the photographic industry. It used to be produced in large quantities in coal gasification plants, where it was removed as an unwanted component from the gas produced. The soil on the sites of such industries is commonly contaminated with cyanide. The threat to human health and the environment posed by such locations greatly depends on the toxicity of cyanide and on its physicochemical behavior, both of which are strongly related to its chemical speciation. The distribution over free cyanide [HCN(aq) and CN-] and iron cyanide complexes [or hexacyanoferrates, e.g., Fe(CN),3and Fe(CN),*-] is especially important. In its free form cyanide can be volatile [HCN(g)], is biodegradable (2, 3), and is much more toxic than cyanide complexed with iron. The complexed forms will interact more with the soil solid phase because of their ionic nature and their ability to precipitate. This may have a great impact on transport rates in soils. Cyanide is mainly disposed of in the form of dissolved iron cyanide complexes or iron cyanide minerals [e.g., 0013-936X/92/0926-0511$03.00/0
Fe,(Fe(CN),),]. In the Netherlands, hundreds of former gasworks sites are contaminated with large amounts of iron cyanide minerals, which cause high concentrations of cyanide in the groundwater. The partition of dissolved cyanide over the free and the complexed form is in the first place determined by the thermodynamic and kinetic stability of iron cyanide complexes. These complexes are generally considered to be extraordinarily stable and are even called kinetically inert ( 4 ) . To be able to determine the cyanide chemically, these complexes have to be broken down. The usual techniques are irradiation with ultraviolet light or a treatment with boiling acid (5-7). Less rigorous, but much slower, ways to decompose these complexes are illumination with visible light or the recently reported microbiologically mediated decomposition (8). However, no data are available about the thermodynamic stability or spontaneous decomposition rates of these complexes under pH and redox conditions found in soils [pH ranging from ca. 3.5 in acidic to ca. 8.5 in alkaline soils and pe ranging from ca. -4 in reduced to ca. 12 in oxidized soils ( 9 ~ .
Available calculations of the thermodynamic stability of iron cyanide complexes in industrial wastewater show that this stability greatly depends on the pH and the total cyanide concentration (IO). Tauchnitz et al. looked at cyanide concentrations of 0.75-6 M, which are much higher than those found in contaminated groundwater. They ignored the effect of the redox potential. Therefore their results cannot be used to represent the speciation of cyanide in soil solutions. Our initial aim was to elucidate the thermodynamic stability of iron cyanide complexes under the pH and redox conditions found in soils and to find out how this stability is affected by changes in these parameters. To do this, we calculated the chemical equilibrium situation for different values of these parameters. But the results of these calculations did not agree with the speciation of cyanide we measured in samples of groundwater from contaminated sites. This suggests that cyanide speciation is not governed by thermodynamic equilibrium but by the kinetics of the slow decomposition of iron cyanide complexes. We decided to pursue this line of research further and to quantify the decomposition rate of iron cyanide complexes as a function of the major soil parameters (Le., the pH, the pe, and the temperature) that determine that rate. Experimental Section Chemical Equilibrium Calculations. A computer model based on MINEQL (11) was used to calculate the cyanide and hexacyanoferrate speciation in aqueous solutions. The input required consists of the following: total amounts or activities of the components involved; relevant
0 1992 American Chemical Society
Environ. Sci. Technol., Vol. 26, No. 3, 1992 511
