Environ. Sei. Technol. 1992,26, 1234-1242
(22) Tang, H.; Eatough, D. J.; Lewis, E. A.; Hansen, L. D.; Gunther, K.; Belnap, D. M.; Crawford, J. Proceedings of the 1989 EPA/A&WMA Symposium on Measurement of Toxic and Related Air Pollutants; Raleigh, NC, May 1989;
Air and Waste Management Association: Pittsburgh, PA, 1989; pp 596-605. (23) Nelson, P. R.; Heavner, D. L.; Oldaker, G. B., I11 Proceedings of the 1990 E P A / A & W M A Symposium on Measurement of Toxic and Reluted Air Pollutants; Raleigh, NC, May 1990; Air and Waste Management Association: Pittsburgh, PA, 1990; pp 550-555. (24) Nelson, P. R.; Ogden, M. W.; Maiolo, K. C.; Heavner, D. L.; Collie, B. B. INDOOR AIR '90-Proceedings of the 5th International Conference on Indoor Air Quality and Climate; Toronto, Ontario, July 29-August 3, 1990; International Conference on Indoor Air Quality and Climate, Inc.: Ottawa, Ontario, Canada, 1990; Vol. 2, pp 367-372. (25) Nelson, P. R.; deBethizy, J. D.; Davis, R. A.; Oldaker, G. B., III Proceedings of the 1991EPA/A&WMA Symposium on Measurement of Toxic and Related Air Pollutants; Raleigh, NC, May 1991; Air and Waste Management Association: Pittsburgh, PA, 1991; pp 449-454. (26) Nelson, P. R.; Ogden, M. W. Proceedings of the 38th ASMS Conference on Mass Spectrometry and Allied Topics; Tucson, AZ, June 3-8, 1990; pp 677-678. (27) Ogden, M. W. In Capillary Chromatography-The Applications; Jennings, W. G., Nikelly, J. G., Eds.; Huethig:
Mamaroneck, NY, 1991; pp 67-82. (28) Jenkins, R. A.; Thompson, C. V.; Higgins, C. E.; Haas, J W., I11 Comparative evaluation of analyticalmethods for the determination of ambient nicotine. Presented et the 40th Tobacco Chemists' Research Conference, Knoxville, TN, October 1986. (29) Thompson,C. V.; Higgins, C. E.; Jenkins, R. A. Compririwn of personal monitoring systems for ambient nicotine Presented at the 41st Tobacco Chemists' Research Con. ference, Greensboro, NC, October 1987. (30) Taylor, J. K. Quality Assurance of Chemical Measure. ments; Lewis: Chelsea, MI, 1987; pp 79-83. (31) Miller, J. C.; Miller, J. N. Statistics for Analytical ('hem. istry, 2nd ed.; Ellis Horwood Limited: Chichester,England, 1988; pp 115-117. (32) Harper, M.; Purnell, C. J. J . Am. Ind. Hyg. Assoc. 1987, 48,214-218. (33) Lewis, R. G.; Mulik, J. D.; Coutant,R. W.; Wooten, \E. W.; McMillin, C. R. Anal. Chem. 1985, 57, 214-219. (34) Feigley, C. E.; Lee, B. M. J . Am. Ind. Hyg. Assoc. 1987,48, 873-876. (35) Perry, R. H.; Green, D. W.; Maloney, J. O., Eds. Perry's Chemical Engineers' Handbook, 6th ed.; McGraw Hill: New York, 1984; pp 3-285-3-287. Received for review October 30, 1991. Revised manuscript re. ceived February 10, 1992. Accepted February 26, 1992
Sorption of Low Molecular Weight Organic Contaminants by Fly Ash: Considerations for the Enhancement of Cutoff Barrier Performance Henry V. Mott*st and Walter J. Weber, Jr.$ Department of Civil Engineering, South Dakota School of Mines and Technology, Rapid City, South Dakota 57701, and Department of Civil and Environmental Engineering, The University of Michigan, Ann Arbor, Michigan 48 109
rn Experimental measurements indicate that class F fly ashes have significant capacity for sorption of low molecular weight organic contaminants from aqueous solution, the capacity being related to the carbon content and to other properties specific to the fly ashes tested. Correlations of Freundlich sorption capacity parameters with the respective octanol-water partition coefficients and the aqueous solubilities of the organic contaminants were highly significant. Transient diffusion experiments verified that incorporation of fly ash into soil-bentonite cutoff barriers can impart significant capacity for contaminant retardation in these barriers. Analysis of the experimental data suggests that the sorption process is rate limited and that assumptions of local equilibrium are not appropriate in the diffusive regime of a soil-bentonite cutoff barrier. Simulations of the performance of a hypothetical soilbentonite cutoff barrier amended with high-carbon fly ash suggest that such amendments can provide significant improvement in performance with respect to retardation of contaminant breakthrough. Introduction It commonly has been assumed that the migration of water-soluble contaminants through soil-bentonite cutoff barriers can be effectively curtailed by restricting the hydraulic conductivity of such barriers to less than cm/s (1-4). Gray and Weber (5), however, suggested that molecular diffusion could be a major transport mechanism 'South Dakota School of Mines and Technology. *The University of Michigan. 1234
Envlron. Scl. Technol., Vol. 26,No. 6, 1992
in such barriers, and Johnson et al. (6)found that diffusion was highly significant in transport of organic contaminants into natural clay beneath a hazardous waste landtill. In an earlier paper (7) we concluded that diffusion could indeed be highly significant as a transport process under conditions of extremely low hydraulic Conductivity. The magnitudes of diffusion coefficients for organic solutes in soil-bentonite cutoff mixtures were found to be reduced from their corresponding free liquid values by factors of approximately one-fourth to two-fifths. Water immobilized by bentonite was found to act essentially as fret. liquid in the diffusion process, and increases in resistance to diffusive transport over free liquid conditions were attributed to decreased open area for diffusion and to tortuosity of the pore matrix. Given this low resistance to diffusive transport, investigation of means for enhancing the retardation capacity of such barriers was rdcom mended. The present work represents the first characterizations of the sorption of specific organic contaminants from aqueous solutions by high-carbon fly ashes, quantifies the sorption of target low molecular weight solutes by such fly ashes, and examines the ramifications such sarptipn would have upon contaminant retardation in soil-bentoa* cutoff barriers. The water adjacent to cutoff barriers can be iii close proximity to organic-phase contaminants, and aqueous concentrations can approach corresponding solubility 1!m its; thus, sorption equilibria were examined over wide concentration ranges with data fitted to the nonlinea Freundlich sorption model: q e = KFC,"
0013-936X/92/0926-1234$03.00/0
(1)
0 1992 American Chemical SocietY
whereqe is the adsorbed-phase concentration (it4 it4-l) at @&brim, KF is the Freundlich unit capacity parameter (MM-'L3" M-"), C, is the solute concentration in the phase ( M L-3)at equilibrium, and n is an exponentialparameter related to the distribution of sorptionenergies. More detailed discussion of the significance of n with regard to distribution-site energies is presented by Weber et al. (8). If the magnitude of n is unity, the may be described in terms of a single distribution coefficient, KD, such that q e = KDce (2) Values of KD have been related to corresponding values of an organic-carbon-normalized partition coefficient, K, though the mass fraction of organic carbon, foc, of soils and sediments such that Koc = K d f o c (3) Numerous correlations between Koc and octanol-water partition coefficients,KOw, and between Koc and aqueous solubilities, S, have been developed for various classes of compounds. Many of these correlations have been compiled and evaluated by Lymann et al. (9). An extension of this approach employed herein for nonlinear sorption involves the correlation of both KF and n with Kow and with S. Combination of the transport processes of diffusion and convection (advection) with the separation process of sorption in a one-dimensional transient model applicable to contaminant migration in soil-bentonite cutoff barriers results in
where t is the porosity of the cutoff barrier, Dei (L2t-l) is the effective diffusivity of solute i in the barrier, u, ( L t-l) is the superficial fluid velocity within the barrier, and ps b the density of the sorbent phase (it4 L-,). If the local equilibrium assumption (LEA) is applied, eq 4 may be more simply stated as
+
where RnI= t p,(l - t)nKFC"-l for a nonlinear sorption relationship described by the Freundlich model. For a linear sorption relationship n is equal to unity and KF is replaced by KD; thus, R1 = t + p,(l - ~ ) K D . If the LEA is not appropriate, the sorption term must $ke a formulation appropriate for sorption rates that are u t e d by one or more mass-transfer resistances. Various of these formulations are described in detail by Weber et al. (IO). The formulation considered here, one of the h p l e s t of these alternatives, is based on the assumptions that the diffusive resistance of an external liquid boundary b e r surrounding each solid sorbent particle controls the rate of solute transfer from the liquid phase to the particle and that the sorption process itself may be described by local equilibrium within the particle
WC, - G,J (6) Where kf ( L t-l) is a mass-transfer coefficient (often defined tu the solute diffusivity divided by the boundary layer thickness), CY, (L2L-3) is the interfacial surface area for pl@s transfer per bulk unit volume of porous medium, C, 18 the solute concentration in the bulk aqueous phase, and %,i is the solute concentration in the aqueous phase in contact with the surface of the sorbent at the exterior of particle. The magnitude of kf may be approximated p,(l- m q , / a t )
for spherical sorbent particles from the following correlation (11, 12) kd/D, 2.0 + 0.6(d~/?)'/~(?/D1)'/~ (7) where d is the characteristic dimension (usually the diameter) of a spherical sorbent particle (L),D, is free liquid diffusivity (L2t-l) of the solute in question, u ( L t-l) is the velocity of the mobile phase relative to the particle, and 7 is the kinematic viscosity of the mobile phase (L2t-l). In a diffusive regime such as that of a soil-bentonite cutoff, the contribution of the second term on the right-hand side of eq 7 is diminished such that the limiting value of the RHS is 2.0, corresponding to diffusion of a single component at steady state outward from a single sphere into an infinite, quiescent medium. Substituting eq 6 into eq 4, rendering the sorption term to be dimensionless with respect to the diffusion term, and considering the limiting case in which q i = 0 (Cs,i= 0) leads to the definition of a Damkohler number, DAD, for a diffusive regime: DAD = kPsd2/D,,i
(8) Other Damkohler numbers (DAc) have been defined for systems in which the convective term of eq 4 is orders of magnitude greater than the diffusive term. It has been found that when the magnitude of DAc is 100 or greater, simulations of solute transport using nonequilibrium formulations of eq 4 agree well with those that employ the LEA (13,14). The mathematics applicable to diffusive and convective systems should behave similarly with respect to the appropriateness of the LEA, thus, if DAD approaches 100 for a diffusive regime, the LEA would likely be appropriate. The objectives of this work were (1) to examine the capacity of class F fly ash for sorption of low molecular weight organic contaminants; (2) to examine the capability of fly ash to retard the transport of organic contaminants in soil-bentonite barriers; (3) to test the appropriateness of the local equilibrium assumption for the description of the sorption process in the diffusive regime of soil-bentonite cutoff barriers; and (4) to examine the effects that incorporation of fly ash into soil-bentonite cutoffs might have on the performance of such barriers with respect to solute breakthrough. Experimental Section
Sorption equilibrium data were obtained by completely mixed batch reactor (CMBR) experiments of the tumbled-bottle type. Fly ash was introduced into individual reactors in dry form and subsequently wetted with a known volume of buffered deionized-distilled water (DDW) to fully saturate the porous fly ash particles prior to the introduction of aqueous solution containing the solute of interest. The DDW contained 0.00307 M NaN, to prevent biological activity and 0.008 73 M NaHCO, as a buffer, such that solutions equilibrated with fly ash were of pH between 7.2 and 7.4. Aqueous solutions containing target solutes were prepared by vigorously stirring mixtures of the respective pure solute and buffered DDW in glass-stoppered flasks for periods of approximately 24 h, which resulted in virtually saturated solutions. Aqueous solutions of desired initial concentration were then prepared from these saturated solutions by dilution with buffered DDW in a 100-cm3glass syringe. Three glass vials fitted with two-piece Teflon-lined screw caps, two of which contained wetted fly ash, were filled from the syringe leaving no head space, sealed immediately, and allowed to equilibrate using gentle end-over-end tumbling for a minimum of 1week. The vials containing fly ash yielded a replicate pair of sorption data points, and the vial conEnviron. Sci. Technol., Vol. 26, No. 6,1992
1235
1/8" BRASS ROD
OLlD BRASS DISC
O L D BRASS DISC
ONFINEMEN SAND edium and coarse)
ERFORATED BRASS
LEXIGLASS DISCS 1/8' BRASS ROD
5 mm I.D. GLASS
Figure 1. Schematic of column for transient diffusion experiments. LIQUID
/--
DAMPENER ---IoN* FLASK
l-w
RETURN LINE TERMINUS \ELEVATION
LEVELS>
n
I
I
-1-1W
I I I
'/ I \
/ I \
I
property
Karn
fly ash Trenton
density: g/cms loss on ignition: % % carbonb specific surface; m2/g raw fired at 600 O C for 24 h
2.34 6.5 4.69
2.29 9.1 6.14
Cob< 2.24 10.3 6.52
1.14
2.65 3.F2 1.08 1.18 Bergstrom (16). Measured as COz recovered during wet corn. bustion. Primary surface area by BET nitrogen adsorption. 0.78
-
soil column. After a predetermined elapsed time, each of the columns was removed from the system, and the con. tents were extruded and sectioned into segments ranging in thickness between 1 and 5 mm. These sections, con. taining both water and soil, were then subjected to liq uid-liquid extraction with hexane over a 24-h contact period. The resulting hexane extracts were assayed to provide time-concentration profiles of total concentration per bulk volume for each column thus treated. Additional details regarding these experimental procedures are given elsewhere (7, 25). The compound used in these experiments was lindane, which is known to degrade abiotically; thus, a fmt-order transformation term was included in the system model. The magnitude of kT was determined on the basis of losses of lindane during quasi-steady state diffusion experiments performed using lindane 8s the transporting solute. Aqueous solutions contained 200 mg/L NaN3as a biological poison; thus, these overall lows were ascribed to abiotic solvolysis (7, 15). Conv7ective transport was absent from the experiments by design. Local equilibrium was assumed as a first cut to analyze the validity of this assumption, and the Freundlicii sorp tion relationship was employed, yielding the following one-dimensional model used for initial attempts to simulate the observed system behavior:
h \
LU
RMP PERISTALTIC
Figure 2. Arrangement of apparatus for transient diffusion experiments.
taining no fly ash was used as a control from which the initial concentrations for the two corresponding sorption points were obtained. The mass of sorbent and volume of aqueous solution contained in each vial were determined gravimetrically. Transient diffusion experiments were conducted to determine the effect of incorporation of fly ash into soilbentonite cutoff mixtures. The device employed in these experiments and the arrangement of the apparatus are shown schematically in Figures 1 and 2, respectively. Experiments were conducted by allowing dissolved solutes to diffuse into the packed columns from the small reservoir located directly below the boundary between the aqueous solution and the column packing. Samples were taken daily from the inlet and outlet manifolds of the solute regeneration and recirculation system to enable definition of a known concentration condition at the boundary of the 1236 Envlron. Scl. Technol., Vol. 26, No. 6, 1992
\
Table I. Properties of Fly Ashes
(114 Dia.)
Samples of fly ash were obtained from the B. C.Cobb, Karn, and Trenton electrical power-generating plants owned and operated by the Consumers Power Co. of Michigan. These fly ashes are derived from bituminous coals mined in the eastern United States and are repre. sentative of ashes from such coals that are burned irl older, non-state-of-the-art electrical-generating plants. The fly ashes were tested for loss on ignition, carbon content, surface area by the BET nitrogen adsorption methcd (both before and after firing at 600 "C for 24 h), and various other properties. The results of these characterizations are shown in Table I. The background soil used in the laboratory to approp imate the native soils used in an optimum soil-bentonlte cutoff mixture was comprised by 77 % silica sand (Agsco Corp., Wheeling, IL, special order), 10.5% silica flour (Agsco 140), and 12.5% kaolinite (Georgia Kaolin). They matwials were chosen for their properties of low organ1c carbon content, high purity, and relatively inert :nine:d surfaces. Grain size distributions for each of the constlt' uents and for the composite background soil are shown in Figure 3. Sodium bentonite (Slurry Ben 90,Aanerlcu Colloid Co.) was incorporated into the background sodat 4 g/100 g of the dry mixture. Other engineering proPAles of this soil mixture are given in a previous paper (7). Seven target solutes [carbon tetrachloride (CTET), trichloroethene (TCE), tetrachloroethene (TTCE),
G 1 e 11. Properties of Target Organic Solutes at 25 "C Kaolinite
Henry's aq constant log solubility WC),atm KOW" (SI,mg/L
diffusivity mol x IO6 radius; cmz/s A
80
I
CTET TCE DCE DCB TCB PCP lindane
1659" 521" 965" 138' 19Oe 0.05e
2.64 2.29 2.60 3.39 3.98 2.39 3.72
785g 1100 1268 79 309 271009 7.28
0.983 1.068 0.961 0.920 0.847 0.930 0.635
3.55 3.49 3.70 3.79 3.98 3.67 4.68
70
3 6ot9
50
L
40
OHansch and Leo (17). bHayduk and Laudie correlation (9). cBmed on LeBas volume (9). "Gosset (18). eEstimated from vapor pressure and solubility data. 'Mackay and Leinonen (19). 6V&eS given a t 20 "C.
b 30 20
-
10
n
Table 111. Columns and Conditions for Solute Assay solute
column/mobile phase
CTET, TCE, 10 ft X 2 mm glass, 6% OV-11 + 4% OV-101 on TTCE Chromosorb W-HP 100/ 120, Ar-CH4 mobile phase DCB, TCB 6 ft X 2 mm glass, 1% OV-351 on Chromosorb W-HP 100/120, Ar-CH4 mobile phase lindane 6 ft X 2 mm glass, 3% OV-1 on Gas Chrom Q 80/100, Ar-CH4 mobile phase PCP 150 mm X 4.6 mm, 5 pm C-18, 80/20 acetonitrilewater mobile phase
\
0 1
1 0
temp, OC detector
0 01
0.001
Grain S i z e (mm)
ECD
Figure 3. Grain size gradations of experimental soils.
ECD
samples of PCP were assayed by reverse-phase, highperformance liquid chromatography. Columns and conditions employed for these assays are given in Table 111.
190
ECD
Results and Discussion Sorption Studies. Karn, Trenton, and Cobb fly ashes
ambient
UV 235 nm
75
75
dichlorobenzene (DCB), 1,2,4-trichlorobenzene (TCB), 4-chlorophenol (PCP), lindane] were chosen to represent several classes of priority pollutants of environmental concern. A listing of these solutes and pertinent properties is given in Table 11. All chemicals used were of reagent grade or better. Aqueous concentrations of CTET, TCE, TTCE, DCB, TCB, and lindane were assayed by gas chromatography with electron capture detection. Liquid-liquid extraction with hexane was used for sample preparation. Aqueous
were tested without modification for sorption of CTET, TCE, T E E , DCB, TCB, PCP, and lindane. Additionally, the fly ashes were tested for sorption of DCB after they had been fired in a muffle furnace at 600 "C for 24 h. The latter experiments showed no detectable sorption of DCB by the ashes after firing, and it was concluded that the fraction lost upon ignition was responsible for virtually all of the sorption capacities of the raw fly ashes. The ratio of carbon to loss on ignition (LOI) was consistent (0.64-0.66) among the fly ashes; thus, the KF values obtained from the sorption experiments were then based on the carbon fraction rather than upon either LO1 or total mass. The resulting parameters obtained by least-squares fitting of sorption data to the logarithmic form of eq 1are listed in Table N. The values of KFfor each solute shown in Table IV were plotted against the respective octanol-
Table IV. Freundlich Sorption Capacity (KF)and Intensity (n ) Parameters for Sorption by High-Carbon Fly Ash solute
fly ash
KFB
95% CIb
n
95% CI*
CV'
CTET
Karn Trenton Cobb Karn Trenton Cobb Karn Trenton Cobb Karn Trenton Cobb Karn Trenton Cobb Karn Trenton Cobb Karn Trenton Cobb
0.348 0.387 1.53 0.658 0.920 1.71 1.37 2.39 4.88 4.02 4.71 6.20 8.66 5.85 14.0 2.95 3.10 8.09 8.05 7.28 19.5
0.160-0.756 0.156-0.961 0.777-3.00 0.400-1.08 0.581-1.46 1.04-2.82 0.808-2.33 1.70-3.36 3.09-7.69 2.82-5.72 3.82-5.80 3.39-11.4 6.67-11.2 4.35-7.87 9.07-21.6 2.01-4.33 2.43-3.96 6.13-10.7 5.02-12.9 4.35-12.2 8.55-44.3
0.609 0.745 0.581 0.477 0.498 0.428 0.466 0.320 0.428 0.322 0.267 0.386 0.256 0.360 0.305 0.267 0.211 0.149 0.348 0.330 0.430
0.356-0.863 0.477-1.01 0.258-0.903 0.309-0.644 0.343-0.653 0.2644.593 0.204-0.729 0.142-0.499 0.170-0.685 0.181-0.463 0.174-0.361 0.172-0.600 0.134-0.378 0.211-0.508 0.180-0.430 0.131-0.404 0.121-0.302 0.049-0.250 0.202-0.494 0.192-0.467 0.252-0.607
161 141 48 145 111 55 27 12 18 20 14 40 18 32 42 14 8 3 65 81 103
TCE TTCE DCB TCB PCP lindane
I 'units of KF correspond to mg/L and mg/g for aqueous- and solid-phase concentrations, respectively. bGuttman and Wilks (20). Coefficient of variation.
Envlron. Sci. Technol., Vol. 26, No. 6, 1992
1237
loo
c
100
KARN +TRENTON
COBB
\
!
\ * 10
KF
KF
1
l !
.1 0
Kow Flgure 4. Relationship between Freundlich K F and octanoVwater partition coefficient for sorption of representative halogenated organic contaminants by Karn, Trenton, and Cobb fly ashes.
Table V. Regressions of KFon Kow and K Fon S significance,* fly ash
regression eqa
R,
KF = 0.0048KoW0.837 0.94 Karn KF = 0 . 0 1 8 7 K o ~ ~ 0.87 .~~ Trenton KF = 0.0468K0W~.~~~ 0.94 Cobb Karn + Trenton KF = 0.0485S-0.521 0.90 Cobb KF = 0.1620S0.474 0.98 Units of S are mol/L.
%
1
5 1 1 0.05
* n - 2 = 4.
water partition coefficients, Kow,in Figure 4 and against the respective aqueous solubilities, S, in Figure 5. It may be noted that the data for PCP are displaced substantially from the trend established by the other solutes. This differing behavior is probably related to its high aqueous solubility (see Table 11) resulting from the presence of the polar OH- functional group. Because of this differing molecular characteristic, the data for this solute are not included in the regressions shown in Figures 4 and 5. The data for CTET also appear to lie somewhat apart from the regressions, suggesting that the behavior of heavily chlorinated single-carbon compounds may be different from that of two-carbon compounds with regard to sorption capacity. Such a hypothesis has yet to be evaluated experimentally; thus, the CTET data are included in the regressions. The resulting regression equations and statistics are listed in Table V. The correlations are highly significant, suggesting that certain physicochemical relationships exist. Of particular interest is the fact that the carbon associated with the respective fly ashes exhibited variations in sorption capacity. The sorptivity ranking was Cobb > Trenton N Karn. Statistical tests indicate that the relationship for the Cobb ash is different at the 95% confidence level from those for the Karn and Trenton ashes, but that the relationships for the two latter ashes are not different. The surface areas associated with the fractions of the fly ashes lost on ignition were determined from the difference between measurements of the surface areas of the raw and fired fly ashes. The surface meas lost on ignition were 6.3,18, and 24 m2/g for Karn, Trenton, Cobb ashes, respectively. The preexponential coefficients shown in 1238 Envlron. Sci. Technol., Vol. 26, No. 6, 1992
,0001
,001
01
.1
1
S (mol/l) Flgure 5. Relationship between Freundlich K F and aqueous solubility for sorption of representative halogenated organic contaminants by Karn, Trenton, and Cobb fly ashes.
Table V are in fact measures of reference sorption ca. pacities of the fly ashes for solutes of KoW equal to unity. These baseline values were divided by the respective surface areas lost on ignition to obtain measures of a baseline specific surface sorptivity. Respective values based on the KF - KOw relationships for Karn, Trenton, and Cobb ashes are 7.6 X 1.0 X and 2.0 X 10”. The three distinct values of this baseline measure of specific sorptivity suggest that sorption-site density is different for the carbon associated with each of the fly ashes. The values of the Freundlich exponential parameter, n, for the three fly ashes and seven solutes were also regressed against both KO, and S. Statistically significant regres. sions were not obtained, although a general trend of lower n value (greater nonlinearity and apparently greater heterogeneity of sorption-site energies) with greater KOW and lower S was observed. The sorption data presented and discussed above indicate clearly that sorption equilibrium is nonlinear over the broad ranges of concentration investigated. Most prior correlations of sorption capacities with the organic cabon associated with natural soils and sediments have been developed for data obtained in narrow and very low concentration ranges, for which phase-distribution reldtionships are often found to be functionally linear. Hence, 8 carbon-normalized distribution coefficient, KD,c, wtis d e fined to compare the sorptivity of fly ash carbo11 and naturally occurring organic carbon: KD,C = KFCen/Ce (10) The value of C, chosen for all solutes investigated was 100 pg/L, and the resulting estimated values of Kc were r e gressed on a log-log basis against both Kow and s in millimoles per liter. The results of these regressions byed on KD,c are compared in Table VI with representatlVe regressions from the literature that are based on K w By comparison of the correlations presented in Table WI,the conclusion may be drawn that at low solute concentrstlom the sorption capacity of the carbon associated with the fly ashes examined in this investigation is at least as great that of organic carbon normally associated with n a t s d y occurring solids.
Table VII. Parameter Values Used in Simulations of Lindane Migration in Transient Diffusion Experiments
Gble VI. Regressions of K c on Kow and K c on S significance, sorbent
regression eqa
Kun and Trenton ash Cobb ash natural sediments (211 Ksn and Trenton ash Cobb ash natural sediments
log Kc = log Kow + 1.17 log Kc = log Kow + 2.15 log Koc = log KO, - 0.21
R,
%
0.84
IC
0.90
lb
0.87 - 0.45 log S log Kc = 5.81 - 0.44 log S 0.94 log Koc = 4.28 - 0.56 log S log K c = 5.65
le lb
(22)
-
parameter
value
source
D,, in confinement, cm'/s De,,in soil-bentonite, cm'/s Freundlich KF for soil" Freundlich n for soil Freundlich KF for fly ash carbon" Freundlich n for fly ash carbon fluid velocity ( U J , cm/s transformation rate coefficient (kT),d-'
1.81 X 10" 2.41 X 10" 3.58 X 0.854 9.84,b 19.47' 0.430
7, 15 7, 15 7, 15 7, 15 this work this work 7, 15 7, 15
0.0 9.02
X
loa
Units are (mg g-' L " mg-"). Most probable value (see Table IV). 'Upper 95% confidence value (see Table IV).
OUnits of S are mmol/L. b n - 2 = 4. cn - 2 = IO.
ExperimentalColumn at 384.3 hr ExperimentalColumn at 194.4 hr Slrnulatlon: Fly Ash Carbon K,= 19.47 Slrnulatlon: Fly Ash Carbon K, 9.84
Slrnulatlon: Fly Ash Carbon K f = 19.47 Slrnulatlon: Fly Ash Carbon K,= 9.84
4\
2000
1000
\LQ , l
0
0
1
3
I
J
"
.
" I
.
0
-m .m
2
0
3
Diffusion Studies. Three columns were packed to a depth of 10.5 cm with the soil-bentonite mixture to which 0.31 g of Cobb fly ash had been added per 100 g whole soil, and the soil-bentonite mixes contained within all columns were then consolidated to 4 psi. A transient diffusion experiment was conducted using lindane as the diffusing solute, which yielded distinct concentration profiles corresponding to three experimental times (one for each column). A time-linearized, fully implicit finite difference approximation of eq 9 was used to generate time-concentration profiles in attempts to reproduce these expermental concentration profiles. Details of the discretization and the FORTRAN computer code used are given elsewhere (15). The value of the effective diffusion coefficient for lindane in a 4% soil-bentonite mixture consolidated to 4 Psi, and the magnitude of the first-order transformation rate coefficient was obtained from previous work (7). Sorption capacity parameters used were those for the soil-bentonite mixture (7, 15) and for Cobb fly ash as reported herein. The most-probable and upper 95% confidence values of KF for fly ash carbon were used to enable consideration of a statistical range of sorption caPacity. A listing of the parameter values used in these Budationsis given in Table VII. The profiles and two elected corresponding simulations for the three experi?ental times are shown in Figures 6-8, respectively. It obvious that these simulations fail to describe the leading %e of the concentration profile. Conversely, at low
3
5
4
DISTANCE FROM INTERFACE (cm)
DISTANCE FROM INTERFACE (cm)
Fbwe 6. Simulated and experimental concentration profiles at 194.4 h foc bansient diffusion experiments using a 4 % soil-bentonite mixture amended with 0.31YO Cobb fly ash.
2
1
Figure 7. Simulated and experimental concentration profiles at 384.3 h for transient diffusion experiments using a 4% soil-bentonite mixture amended with 0.31 % Cobb fly ash.
Experimental Column at 774.6 hr Slmulatlon: Fly Ash Carbon K,= 19.47
z
Slmulatlon: Fly Ash Carbon KF= 9.84
4000
0
-I-
2000
??
e-
0
I
2
3
DISTANCE
4
5
6
7
8
9
10
FROM INTERFACE (cm)
Figure 8. Simulated and experimental concentration profiles at 774.6 h for transient diffusion experiments using a 4% soil-bentonite mixture amended with 0.31 % Cobb fly ash.
penetrations, where the profile is well-developed,conditions approach equilibrium, and the simulations and exEnviron. Sci. Technol., Vol. 26, No. 6, 1992
1239
perimental profiles agree reasonably well. On the basis of an effective diameter, dlo,of 0.004 mm (16) of the Cobb fly ash particles and an estimation of kf using the limiting diffusive case for eq 7, DAD was computed for conditions corresponding to those in the vicinity of the leading edge of the concentration profile (qi = 0) and found to be 0.04, well below the value of 100 suggested as appropriate for use of the LEA ( 1 3 , 1 4 ) . Failure of the LEA to describe these experiments satisfactorily would then be expected. The magnitude of a, is directly related to the fly ash content of the soil-bentonite mixture, and DAD is thus directly proportional to a,. Consequently, an increase in the fly ash content of the mixture will result in a corresponding increase in DAD. Hydraulic conductivity experiments conducted by Bergstrom (16) suggest that a fly ash content as high as 40% could be used in soil-bentonite cutoffs without imparting significant adverse effects upon hydraulic conductivity. The value of DAD for such a mixture under conditions similar to those of the transient diffusion experiment was calculated to be approximately 4.0, which also is well below the value of 100 suggested as the minimum value necessary for validity of the local equilibrium assumption. The conclusion to be drawn then is that accurate predictions of solute transport in soilbentonite cutoffs modified by addition of fly ash or other similar materials must consider both nonequilibrium and nonlinear conditions in modeling the sorption term. Accurate long-term predictions of solute migration that would consider these mathematical complexities can be accomplished only after an efficient numerical approximation is assembled. Such an approximation was beyond the scope of this investigation. Simulations of Cutoff Barrier Performance. Two extremes envelop the realm of potential sorption interactions that may occur within a soil-bentonite barrier. These are sorption capacities sufficiently low to be negligible and sorption that may be described by the LEA. Simulations were performed for a hypothetical barrier considering these two conditions for the purpose of delineating the behavior of the two bounding cases. The thickness of the barrier was assumed to be 1m, the areal extent of the containment was assumed to be 10 acres, and the cutoff was assumed to extend 50 f t below an unconfined water table. The assumed shape of the containment was a cylindrical shell, and due to the large diameter and corresponding large radius of curvature, the barrier was idealized as a planar sheet of infinite areal extent. The hypothetical barrier was assumed to be comprised by 4 g of bentonite and 40 g of Cobb fly ash per 100 g of whole soil. Local equilibrium and, for ease of computation, linear sorption relationships were assumed. The solute concentration at the interior face of the barrier, Co, was considered constant at 1mg/L, and the exterior face of the barrier was considered to be continuously flushed, such that the solute concentration at this position was maintained effectively at zero. The analytic solution for the solute flux, Pi,( M P t-l), at the exterior face of the barrier was adapted for the applicable form of eq 5 from Crank (23)considering convection to be negligible:
where TB is the barrier thickness (1 m); DR,i = D , J R , ; C(z,t=O) 0; C(z=O,t) = Co; C(z=TB,t)= 0; and all other terms are as previously defined. The target solute considered was CTET, with the effective diffusion coefficient estimated using the relationship Dei = t4I3Dli(7) based on 1240 Environ. Sci. Technoi., Voi. 26, No. 6, 1992
-
0.6 +
w/o fly ash w/40% Cobb fly ash
05-
ELAPSED TIME (yr)
Flgure 9. Simulations of diffusive flux of carbon tetrachloride ihrough hypothetical soil-bentonite cutoff barriers with and without amendment by fly ash.
a porosity, E , of 0.48. The partition coefficient for the cme of negligible sorption was assumed to be zero, and that for the sorption case was estimated using eq 10 with Ce = 1 mg/L, KF = 1.53, and n = 0.581 (see Table IV). The total solute migration through the barrier per unit area, Qt (M L-'),through the barrier is the integral over time of the flux which is readily approximated numerically ali
1 t
Qt =
FO,,
0
t
dt
= CFo,,At
(12)
0
Plob of F",,versus time for these two hypothetical limiting cases, a barrier with fly ash and a barrier that is virtually devoid of sorption capacity, are shown in Figure 9, and corresponding plots of Qt versus time for these two cases are shown in Figure 10. For the case assuming no sorption capacity, breakthrough is predicted to occur within :i years, and near-steady-state flux rates are predicted to obtain within 12 years. Conversely, for the case assuming amendment of the barrier with fly ash, solute breakthrough is predicted to occur in approximately 30 years, and near-steady-state flux rates are predicted to obtain within approximately 220 years. This retardation of solute breakthrough is in fact only a delay of the inevitable. It may be observed from Figures 9 and 10 that both cases approach identical steady-state flux rates. A delay of 100 years means little to future generations given the persistence of certain organic contaminants in subsurface environments. However, as current design specifications call for materials that result in barriers that are virtudY devoid of sorption capacity, such a delay reprebents a monumental improvement over the expected performance of in-place barriers. In consideration of the discussion given above regarding nonequilibrium and nonlinedY, these predictions are presented solely as bounding calCUlations. The true behavior of any given barrier will fall between the two limiting cases considered here. To provide for proper evaluation of mitigative strategies, it is im perative that methods be developed to more accmtely predict solute transport in the sorption-affected, diffusive regime within these barriers. Moreover, since both solUte breakthrough time and ultimate contaminant flux rates afe highly dependent upon the effective diffusion coefficient, means of enhancing the resistance of soil-bentonlte
+
.-
w/oflyash w/40% Cobbflyash
/+
a P
E
2
c
s @ c
-
0.0
0
100
200
300
400
500
is evidence that other properties specific to individual fly ashes may also be important. The results of transient diffusion experiments bear out the contaminant retardation benefits that can accrue from incorporation of high-carbon fly ash into soil-bentonite cutoffs. Attempts to simulate diffusive transport behavior using the assumption of local equilibrium indicate that such an assumption is not appropriate for the conditions examined and that both nonlinear sorption equilibria and mass-transfer limited sorption rates must be considered. Analyses of the experimental behavior using a diffusive Damkohler number suggest that conditions within soilbentonite cutoffs are well-removed from those for which assumptions of local equilibrium may be applied with satisfactory accuracy. Simulations based on a hypothetical barrier performed for two bounding cases suggest that solute migration through soil-bentonite barriers by molecular diffusion can be significant and that the addition of a sorbent phase such as fly ash to soil-bentonite mixtures can markedly improve the performance of such barriers with respect to retardation of organic contaminant migration.
ELAPSED TIME (YR)
Simulations of cumulative migration of carbon tetrachloride through hypothetical soil-bentonite cutoff barriers with and without amendment by fly ash. 10.
barriers to diffusive transport should be sought if these barriers are to be viable alternatives for mitigation of subsurface contamination. It is noted that the work discussed herein considered only single-solute cases. The chromatographic effects operative in multisolute systems could contribute significantly to solute transport. To provide bases for accurate long-term predictions of barrier performance, these effects should be also evaluated with respect to diffusion and sorption within these barriers. The potential negative side of the use of high-carbon fly ash for retardation of organic solute migration is the possibility of leaching of trace metals from the fly ashes and subsequent contamination of subsurface environments by these trace metals. Cations or cationic complexes of trace metals should be well-attenuated within the barrier by the natural cation-exchange capacity of the bentonite (24-27). Conversely, the anionic complexes of trace metals such as As, Se, and Cr, present in most fly ashes, would be of prime concern as most soils, and particularly bentonite, have little natural anion-exchange capacity under conditions prevalent in most soil-bentonite barriers (26-28).The potential for leaching of As, Se, and Cr from high-carbon fly ashes warrants consideration in the evaluation of the suitability of these sorbents for amendment Of soil-bentonite or other earthen barriers. Conclusions
The capacity of residual carbon typically associated with Class F fly ashes from older power plants for sorption of representative halogenated organic contaminants was found to be significant and at least equivalent to that of the organic carbon typically associated with naturally ?curring soils and sediments. Sorption equilibria exhiblted nonlinear behavior over the concentration ranges studied and were found to be described well by the h n d l i c h isotherm model. Freundlich sorption capacity fabrs, KF,were found to correlate well on a log-log basis with the octanol-water partition coefficients and aqueous solubilities of the organic compounds studied. The heudlich exponential parameter, n, did not. The sorption capacities of the fly ashes tested appear to depend 81gnifcantlyon the carbon content of the fly ash, but there
Acknowledgments
We particularly acknowledge two colleagues of the Department of Civil and Environmental Engineering at The University of Michigan for their input to this work: Professor Donald H. Gray for his advice regarding soil properties and characteristics and Professor Will Hansen for his assistance in the determination of specific surface areas. Registry NO, CTET, 56-23-5;TCE, 79-01-6;"M'CE, 127-18-4; DCB, 106-46-7;TCB, 120-82-1;PCP, 106-48-9;C, 7440-44-0; 1octanol, 111-87-5; lindane, 58-89-9. Literature Cited (1) D'Appolonia, D. J. ASCE J . Geotech. Eng. Diu. 1980,106, 399. (2) Schulze, D.; Barvenik, M.; Ayres, J. Proceedings of the
Fourth National Symposium and Exposition on Aquifer Restoration and Ground Water Monitoring, Columbus, OH, May 1984. (3) Jepson, C. P. Pollut. Eng. 1984, 16 (Apr),50. (4) White, L. A.; Daagupta, A.; Coia, M. F. J. Hazard. Mater. 1987, 14, 39. (5) Gray, D. H.; Weber, W. J., Jr. Proceedings of the Seventh
Annual Madison Waste Conference, University of Wisconsin-Extension,Madison, WI, Sept 1984. (6) Johnson, R. L.; Cherry, J. A.; Pankow, J. F. Enuiron. Sci. Technol. 1989, 23, 340. (7) Mott, H. V.; Weber, W. J., Jr. Enuiron. Sci. Technol. 1991, 25 (lo), 1708. (8) Weber, W. J., Jr.; McGinley, P. M.; Katz, L. E. Presented
at the 65th Colloid and Surface Science Symposium,American Chemical Society, Norman, OK, June 1991. (9) Lymann, W. J.; Reehl, W. F.; Rosenblatt, D. H. Handbook of Chemical Property Estimation Methods; McGraw-Hill New York, 1982; Chapters 4, 17. (10) Weber, W. J., Jr.; Katz, L. E.; McGinley,P. M. Water Res. 1991, 25 (5), 499. (11) Bennett, C. 0.; Myers, J. E. Momentum, Heat and Mass Transfer, 3rd ed.; McGraw-Hill: New York, 1982; p 590. (12) Cussler, E. L. Diffusion: Mass Transfer in Fluid S y s t e m ; Cambridge University Press: New York, 1984; p 231. (13) Jennings,A. A.; Kirkner, D. J. ASCE J. Hyd. Eng. Diu. 1984, 110 (12), 1700. (14) Bahr, J. M.; Rubin, J. Water Resour. Res. 1987,23 (3), 438. (15) Mott, H. V. Ph.D. Dissertation, The University of Michigan, 1989. (16) Bergstrom, W. R. Ph.D. Dissertation, The University of Michigan, 1989.
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Environ. Sci. Technol. 1992, 26, 1242-1248
(17) Hansch, C.; Leo, A. Substituent Constants for Correlation Analysis in Chemistry and Biology; Wiley New York, 1979; pp 172-330. (18) Gossett, J. M. Environ. Sci. Technol. 1987, 21 (2), 202. (19) MacKay, D.; Leinonen, P. J. Environ. Sci. Technol. 1975, 9 (13), 1178. (20) Guttman, I.; Wilks, S. Introductory Engineering Statistics; Wiley: New York, 1965. (21) Karickhoff, S. W.; Brown, D. S.; Scott, T. A. Water Res. 1979, 13, 241. (22) Chiou, C. T.; Peters, L. J.; Freed, V. H. Science 1979,206 (Nov), 831. (23) Crank, J. The Mathematics of Diffusion, 2nd ed.; Clarendon Press: Oxford, U.K.; 1975; pp 49-52. (24) Griffin, R. A.; Shimp, N. F. Environ. Sci. Technol. 1976, 10 (6), 1256. (25) Griffin, R. A; Shimp, N. F.; Steele, J. D.; Ruch, R. R.; White, W. A.; Hughes, G. M. Environ. Sci. Technol. 1976,lO (6), 1262.
(26) Bohn, H. L.; McNeal, B. L.; O'Connor, G. A. Soil Chemistry; Wiley: New York, 1979; pp 141-192. (27) Sposito, G. The Surface Chemistry of Soils; Oxford University Press: New York, 1984; pp 1-35. (28) Griffin, R. A.; Au, A. K.; Frat, R. R. J. Environ. Sci. Health 1977, A12 (8), 431.
Received for review August 5,1991. Revised manuscript received January 28,1992. Accepted February 18,1992. Partial support for this work was provided, in sequence, by Research Grant R811570-01-0 from the U.S. Environmental Protection Agency, by a University of Michigan Rackham Research Fellowship award to H.V.M.,and by Research Grant l-P42-ES04911-01 from the National Institutes of Environmental and Health Sciences. Neither the Environmental Protection Agency nor the National Institutes of Environmental and Health Sciences has reLiewed this article, and the opinions and conclusions set forth herein are not necessarily those held by either agency.
Lead Bioavailability: Dissolution Kinetics under Simulated Gastric Conditions Michael V. Ruby,'vt Andy
J. Houston Kempton,t John W. Drexler,f and Paul D. Bergstroms
PTI Environmental Services, 2995 Baseline Road, Suite 202, Boulder, Colorado 80303,Department of Geological Sciences,
University of Colorado at Boulder, Boulder, Colorado 80309,and Atlantic Richfield Company, 555 17th Street, Denver, Colorado 80202 The lower bioavailability of lead from mining sites compared to urban environments is due partially to the relative solubility of Pb-bearing phases in the respective mineral assemblages and to kinetic limitations relative to the residence time of soil in the gastrointestinal (GI) tract. In this study, P b dissolution kinetics for mine-waste-impacted soil and for pure anglesite were diffusion controlled. Dissolution rates over stomach residence time (2 h) were 0.49 and 0.07 [(mg of Pb/L of solution)/(mg of solid Pb)(h)] for PbS04 and test soil, respectively, and were linearly dependent on HC1 concentration, whereas particle size did not affect dissolution rates except at mean diameters of PbS04 > test soil. Dissolution of only 4% C P b from the test soil after 2 h demonstrates that kinetics of Pb-bearing minerals in mine-waste-impacted soils is an important factor controlling P b solubility in the GI tract.
Introduction The factors controlling anthropogenic lead bioavailability have been investigated by numerous authors (1-8). However, only recently have investigations focused on the geochemical controls of P b bioavailability and the mechanisms controlling P b dissolution from soil particles in the gastrointestinal (GI) tract, such as rinding by precipitation products, dissolution kinetics, and encapsulation by alteration products and inert matrices (e.g., silicates) (9, 10). Dissolution kinetics of an ingested solid Pb-bearing phase chsracteristic of a mining site are affected by both Pb mineralogy and particle size distribution. For soluble P b salts such as Pb(OAc)z,which is often used in toxicological P b studies because it dissolves rapidly (4, 8, 11), the concentration of P b in solution is controlled by the available mass of salt. However, for anglesite (PbS04), which may control P b dissolution from soils impacted by PTI Environmental Services.
* University of Colorado at Boulder. 5 Atlantic
1242
Richfield Company.
Environ. Sci. Technol., Vol. 26, No. 6, 1992
mine waste originating from the reduced zone of the ore body, Pb solubilization is controlled by kinetic constraints, resulting in lower bioavailability relative to Pb(OAe1, (12, 13).
This paper describes in vitro P b dissolution kinetics using both mine-waste-impacted soil and pure, crystalline anglesite to estimate the solubility of Pb-bearing solids during passage through the stomach. The effects of particle size and pH were examined, and kinetic modeLs were used to evaluate the mechanism of P b dissolution. The equilibrium geochemical speciation model MINTEQAB (14) was evaluated to determine if it is capable of predicting realistic aqueous P b concentrations for a nonequilibrium system. Experiments such as these, which examine the mechanistic framework for Pb bioavailability, are necessary to evaluate or construct a predictive model for P b bioavailability from soil. Although the subject of anglesite dissolution kinetics has been studied previously under laboratory conditions Us), there have been no previous investigations under physio. logical conditions, or as a function of particle size and pH. An understanding of the kinetic constraints that limit bioavailability of Pb-bearing mine waste is valuable because human health risk assessments used to define cleanup standards at mine waste sites are particvldY sensitive to this parameter.
Soil Chemistry Oxidation reactions on mineral surfaces result in Brmoring of the primary mineral grain by a secondary reaction product. For example, the weathering of galena (PbS) in the acidic environment generated by pyrite dis. solution in mine-waste rock (16) results in a rind of anglesite around a galena core (17). Ingestion of soil mixed with mine waste that bears anglesite results in a dissolution reaction with HC1 in the stomach to form predominantly PbCl+: PbS04(s) + HCl(aq)
0013-936X/92/0926-1242$03.00/0
F?
PbCl+(aq) + HS04-(aq) (l)
0 1992 American Chemical Soclew
