Environ. Sci. Technol. 1990, 24, 1349-1354
Effect of Desorption and Intraparticle Mass Transfer on the Aerobic Biomineralization of a-Hexachlorocyclohexane in a Contaminated Calcareous Soil Huub H. M. Rljnaarts, And& Bachmann,+Johannes C. Jumelet, and Alexander J. B. Zehnder'
Department of Microbiology, Wageningen Agricultural University, 6703 CT Wageningen, The Netherlands
A laboratory study was performed to investigate the effects of desorption from soil aggregates on the biomineralization kinetics of a-hexachlorocyclohexane(a-HCH). Desorption and biodegradation of a-HCH in mixed soil suspensions at 20 "C were shown to be controlled by intraparticle mass-transfer processes. Two models were applied to the desorption and biodegradation kinetic data, a first-order model (FOM) and a sorption-retarded radial diffusion model (RDM). Only the RDM could explain the effect of aggregate size on HCH desorption and bioconversion rate. It also yielded the best fit to the desorption kinetic data. Estimated effective diffusivities were in the order of 5 X lo-'' m2/s. Biodegradation kinetics of a-HCH could only be described by the RDM by assuming that microorganisms could penetrate the inner parts of the aggregates. Introduction Isomers of hexachlorocyclohexane(HCH) are known to be rather persistent xenobiotic compounds in the natural environment (1-4). 0-and 6-HCH have been shown to be extremely recalcitrant compounds due to their stable molecular structure (3,4). Lindane (or y H C H ) and the a-HCH isomer are biodegradable. Their bioconversion rates have been reported to be affected by their own concentration and a number of environmental factors, such as redox condition, temperature, nutrient limitation, and presence of auxiliary organic growth substrates ( 4 , 5 ) . In previous studies it was found that in aerobic mixed soil slurries a-HCH was completely mineralized. The highest rates were measured in the absence of auxiliary organic growth substrates and at a temperature between 15 and 30 "C. a-HCH mineralization was stoichiometrically paralleled by chloride increase ( 4 , 5 ) . Studies with pure microbial cultures isolated from this contaminated soil indicated that 100% of the chlorine and 50% of the organic carbon were released as chloride and C02,respectively (6). Besides environmental factors, sorption processes have also been shown to influence the biodegradation of various organic chemicals in aqueous systems with solid particles (7, 8). As a result of sorption both an increase and a decrease of bioconversion rates are possible (9-11). Sorption often has been found to stimulate biodegradation when the compound to be degraded or its metabolites were toxic to the microorganisms. In this case sorption is responsible for a lower free concentration of the toxic chemicals (11,12). More often, however, sorption has been reported to decrease bioconversion rates (12,13). If microorganisms can only utilize dissolved substrates, as has been shown by Ogram et al. (14) and Chakravatry (1.9, sorption can retard bioconversion in two ways: either by simply decreasing the aqueous concentrations to very low concentration levels, which may even become lower than threshold values for biodegradation (16), or by limiting Present address: MBT Umwelttechnik AG, Vulkanstrasse 110, CH-8048 Zurich, Switzerland. 0013-936X/90/0924-1349$02.50/0
high potential biodegradation rates by small desorption velocities resulting in desorption-controlled bioconversion (12, 17). (De)sorption rates depend on solid- and liquid-phase concentrations. They can be controlled by surface chemical reaction rates (18,19), by the rate of permeation into or out of the organic matter fraction (20,21), or by diffusion through a nonturbulent liquid layer around the solid-phase particles (22-26). For porous adsorbents such as activated carbon, clay, sediment, and soil aggregates, intraaggregate diffusion processes can in addition contribute to a great extent to the retardation of solute transport between bulk liquid and sorption site (22-27). In batch systems mixing intensity is also an important factor. Higher mixing speeds result in enhanced external mass transfer and may even lead to diminished internal diffusion resistances due to a breakup of aggregates (28,29). Diffusion rates strongly depend on the average solute transport distance and the concentration gradient. Therefore, microorganisms situated close to surfaces of porous systems onto which metabolizable substrates are adsorbed can additionally increase the diffusion rates in two ways: The microbial removal of the dissolved part of the substrate will steepen the concentration gradient, and the vicinity of the microorganisms to the adsorption site will reduce the diffusion distances (30). As a consequence, diffusion rates in nonsterile systems are faster as compared to those in sterile systems (31, 32). In the following, data are presented on the kinetics of desorption and biomineralization of a-HCH in suspensions (slurries) made from a contaminated soil. It is shown that in mixed suspension systems intraparticle mass transfer controls the rates of desorption and biomineralization of a-HCH. Materials and Methods
Soil Sampling. Samples of soil were collected at a contaminated site in Hengelo, The Netherlands. Sampling depth was between surface and 50-cm depth. The samples were sieved (2-mm mesh), homogenized, and stored in individual batches at -20 OC until usage, as described earlier (4). Analytical Methods. The concentration of HCH was determined by capillary gas chromatography. HCH was extracted with an acetonitrile-water mixture ( 4 ) . Soil contaminants other than HCH and the metabolites produced during incubation were analyzed by reverse-phase high-pressure liquid chromatography (HPLC) and gas chromatography coupled to mass spectrometry (GC/MS) (4). Oxygen was quantified by gas chromatography (Becker Model 406). The biochemical oxygen demand (BOD) assay was done in a Sapromat (33). a-HCH biodegradation performance was monitored by measurement of the release of free chloride ( 4 ) . Free chloride was determined by potentiometric titration using a Marius Micro-chloro-counter (Marius, Utrecht, Netherlands). Physical and chemical characteristics of the soil were de-
0 1990 American Chemical Society
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termined according to the methods described by Houba (34). Media and Chemicals. The mineral salts medium for the experiments contained, per liter: KH,PO,, 170 mg; K,HPO,, 570 mg; Na2HP04,668 mg; NaHC03, 20 mg; FeC13.6H,0, 2.03 mg; NH4N03,185.5 mg; MgS04, 22.5 mg; CaCl,, 27.5 mg. HCH isomers were obtained from Schmidt B.V., Amsterdam, NL. All chemicals used were reagent grade and were obtained from commercial sources. Contaminated Soil. The contaminated soil used was characterized by a rather high pH of -7.7, a low soil organic content of 1.1% (w/w), and a Kjeldhal nitrogen content of -0.05% (w/w). Total P, K, Na, Ca, and Fe were 0.2, 2.6,0.42, 16.1, and 2.3 g/kg, respectively. The soil BOD was below detection limit (10 mg of O2L-l-day-'). From a total of lo8 microorganisms (measured by direct counts), a t least lo3 aerobic a-HCH-degrading microorganisms (measured by the most probable number technique) per gram of dry soil were present in the contaminated soil ( 4 ) . About 60% of the soil's constituents was sandy material. The original soil was amended with 38% (w/w) lime by the chemical company while disposing their HCH wastes -20 years ago. The solid-phase density (p,) was 2.4 g/cm3. Two different soil samples were used in this study, one (sample A) containing 240 f 50 mg/kg a-HCH and the other (sample B) containing 140 f 30 mg/kg a-HCH. Other concentrations and physical/ chemical characteristics were determined to be equal for both samples. Besides a-HCH, 0-HCH was found to be a major pollutant (220 f 25 mg/kg). y-and 6-HCH were also detected but at much smaller levels, 22 and 13 mg/kg, respectively. The soil was found to be contaminated with various chlorinated benzenes and two dichlorophenol isomers as well. The contamination levels of these aromatic compounds ranged from 0.3 to 4.3 mg/kg ( 4 ) , which is relatively low compared to HCH. Soil Suspension Systems. If not stated otherwise, soil slurries contained 100 g of suspended solids (dry weight)/L of mineral medium. The suspensions were either mixed on an end-over-end (e\e) mixer (8 rpm, amplitude 20 cm) or stirred with a magnetic stir bar. Experiments with the e\e mixed suspensions were conducted with 250- (bioconversion) and 500-mL (desorption) glass bottles containing 115 and 350 mL of slurry, respectively. Experiments with the stirred system were performed with 2-L flat-bottom flasks containing 700 mL of suspension. Aggregate Size Distribution in the soil slurry systems was determined after 1 h of stirring or e\e mixing. A few slurry drops were placed on a glass slide, spread, and air-dried. Six samples were analyzed for each slurry system. A t 50 randomly chosen points of each sample the aggregate size was determined and thereafter classified according to the international standard classes of particle size (35). From the aggregate size distributions (Figure 1) for the two types of suspension systems the numbermean particle diameter (d,) was determined by assuming a log-normal distribution. Assuming spherical particles, the volumetric-mean particle diameter (d,) can be obtained by using the relationship log d, = log d, + 2.3 log u in which ug,, is the geometric standard deviation o f i h e number-mean diameter. For the e\e mixed and the stirred suspensions the values for d, were 182 and 122 pm, respectively. Desorption Studies. Desorption studies were conducted at a temperature of 21 f 2 "C. Various concentrations of sorbed a-HCH were obtained by preincubating an end-over-end (e\e) mixed nonsterile soil suspension. After a certain amount of a-HCH was degraded, sodium 1350
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end -over-end mixed d, = L1 u m d, 182 um 02
-
+
d,: 26" d,
0
122 pm
50 100 250 500 loo0 aggregate size l p m )
10
Flgure 1. Aggregate size distribution of the soil suspensions. d , and d , are the number-mean and the volumetric-mean aggregate size, respectively.
azide was applied in a concentration of 5 mM to stop microbial activity. Sorbed concentrations a t that point were calculated from the total amount of a-HCH present at the start minus the amount degraded as calculated from the chloride release. Samples of -8 mL of soil suspension were taken after certain time intervals. These samples were centrifuged in stainless steel tubes for 15 min at 48000g and the supernatants were analyzed by GC and HPLC. Biodegradation Studies. In both suspension systems biodegradation was studied a t 21 f 2 "C, if not stated otherwise. Aerobiosis was assured by periodically determining the oxygen content of the gas phase, which was flushed with air for 10 min at a flow rate of 200-300 mL/min whenever necessary. In all the slurry experiments bioconversion was monitored by measuring the release of free chloride. Chloride release was already shown to be correlated with HCH biodegradation (4). Killed controls for all the biodegradation experiments were prepared by y radiation of the incubation mixture of 2.5 Mrad (Gammaster, B. V., Ede, NL) followed by pasteurization for 30 min a t 70 "C. As reported earlier ( 4 ) , the contaminants in the soil did not inhibit aerobic respiration of the autochthonous microbial populations already present in the soil. Analytical Precision. All data reported are based on duplicates. Standard errors were within 15% of the values determined by GC analyses. The HCH detection limit in the contaminated soil suspension was -5 mg/kg of dry soil and 10 pg/L when HCH was solubilized in water. Model Fitting and Data Analyses. FORTRAN algorithms were written describing the analytical solutions of the differential equations of the models used. These algorithms (available on request) were linked as a subroutine to the BMDP AR nonlinear regression (NLR) procedure (36). Also, in some cases regular linear regression procedures were used that did not allow for an intercept (LRNI).
-
Results and Discussion 1. Desorption and Intraparticle Mass Transfer. The suspensions with various initial sorbed HCH concentrations were incubated for 24 (stirred) or 48 h (e\e mixed). The distribution coefficient, KD,was determined
~~
Table I. Results of Fitting the Desorption Kinetic Models to the Data by Using the Nonlinear Regression (NLR) Procedure" A. First-Order Model
suspension system stirred end-over-end mixed
kd,
CW
day-' 0.42 (0.04) 0.05 (0.01)
mg/L
Ppsudo
0.94 0.70b
0.85
B. Sorption-Retarded Radial Diffusion Model Concentration in solution (mg ("1
Figure 2. Desorption isotherms for a-hexachlorocyclohexane in end-over-end mixed (E) and stirred (S) soil suspensions.
-
-F T
I
c
2 3
?
.-
1.2
1
08-
5
+
suspension system
Deff,
stirred
Ce
9
10-l~ mz/s
mg/L
rzwud0
6.4 (1.3) 0.7
2.3 (0.3) 4.8 (3.5) 1.3c
0.97
end-over-end 0.92 (0.1) mixed end-over-end 4.1 0.97 mixed (0.4) ONumbers in parentheses are the standard deviations. bCe is calculated by using the measured KD from the stirred system (179 cm3/g). 'C, is calculated by using the calculated KD from the stirred system (94 cm3/a).
S: SO:240mg.kg" 0
.
!F c
J Y I 1 8 radial diffusion
1.0
04 0
4
8
12
16
20
-0-
' F 48
24'
time (hours)
Flgure 3. Observed desorption kinetic data of a-hexachlorocyclohexane in end-over-end mixed (E) and stirred (S) soil suspensions.
from linear isotherms (Figure 2). Linear regression analyses (LRNI) yielded estimated values for KD of 251 (P = 0.987) and 179 cm3/g (r2= 0.924),respectively. The system dependence of the distribution coefficient indicates that, at least in the e\e mixed systems, equilibrium had not been reached yet after 48 h. Desorption kinetic data for the sterile e\e mixed and stirred suspension system are given in Figure 3. Two kinetic models have been fitted to those data: a first-order and a radial diffusion model. The first-order model is derived from Langmuir's isotherm. Assuming a linear isotherm, this model is described by the following equation (18, 19): dS/dt = kd(KDc - 8) (1) where S is sorbed concentration (g/g), C is the solution concentration (g/cm3), t is the time (day), and k d is the desorption rate constant (day-l). Assuming that in the stirred system equilibrium has been achieved after 24 h, KD is set to 179 cm3/g. This implicates that for the e\e mixed suspension at t = 48 h only 60% of the equilibrium concentration was reached. By use of a KDof 179 cm3/g and the nonlinear regression (NLR) procedure, 322s of 0.42 and 0.05 day-' were obtained for the stirred and eye mixed system, respectively. Fitting results are shown in Figure 4A and Table IA. The first-order model fits well the desorption data from the stirred system. However, the relatively fast desorption during the first 8 h in the e\e mixed system is not well described by the model. Surface chemical reaction rates are in general much higher than the observed desorption rates (27, 28). Therefore, two
time (hours)
Figure 4. Observed and predicted (lines) desorption kinetics of ahexachlorocyclohexanein end-over-end mixed (E) and stirred (S) soil suspensions. The models used are a first-order model (A) and a sorption-retarded radial diffusion model (B).
mechanisms can be responsible for the slow desorption: diffusion through a nonturbulent liquid layer around the soil particles or intraparticle mass transfer. If external stationary diffusion controls desorption, k d can be expressed as GD,/d,L, in which D, is the diffusivity for the sorbate in the liquid phase (for HCH this is 1 X m2/s-l), d, is the volumetric-mean diameter of the aggregates, and L is the thickness of the diffusion liquid layer. However, this relationship does not seem to hold for the systems studied here [kd,&red/kd,e\emixed = 8, while d,,e\emixed/dv,stin~ = 1.5 and calculated values of L are b o high ( > l o m) to be realistic]. Therefore, not external diffusion but intraparticle mass transfer must be the cause of the slow desorption observed. Besides being a poor fit of the e\e mixed system, the first-order model can only be used empirically to describe intraparticle mass transfer; it cannot explain why desorption kinetics for the e\e mixed system were slower than for the stirred system. A model that describes intraparticle mass transfer in a more mechanistic way is the sorption-retarded radial diffusion model developed by Wu and Gschwend (27). This model can explain the difference between the two suspension systems because it is based on the aggregate size as the diffusion domain. It is described by the equation Environ. Sci. Technol., Vol. 24, NO. 9, 1990
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- 10 D1
1
- 4.0
with (3)
where S ( r ) is the local total volumetric concentration in porous sorbent (g/cm3), r is the radial aggregate distance (m), Deffis the effective intraparticle diffusivity (m2/s),n is the porosity of the sorbent (m3/m3),7 is the tortuosity factor, and ps is the solid-phase density (g/cm3). Wu and Gschwend applied eq 2 and 3 successfully to suspensions of soil aggregates and natural sediments by assuming instantaneous equilibrium inside aggregates, a linear sorption isotherm, and a sufficiently turbulent bulk fluid ensuring that sorptive exchange is not limited by transport through an exterior boundary layer and that f ( n , r )= n' with 1 I i I 2, to be equal to n, thus i = 1 (27). An analytical solution of eq 2 is available for systems homogeneous in aggregate size and a well-mixed exterior solution volume (37). Cooney and Adesanya (38) showed that this analytical solution is only valid for a mix of particles having a particle size distribution that spans about or less than 1 order of magnitude. In both suspension systems more than 90% of the soil aggregates were within the range of 5-250 gm (Figure 1). Although this range was a little too wide, this analytical solution of eq 2 was applied to the desorption kinetic data by using the NLR procedure and by assuming that all particles had the same size as the volumetric-mean particle size (dJ. Both Deff and the equilibrium concentration in the liquid phase (C,) were taken as independent model parameters since after 48 h equilibrium has not been reached in the e\e mixed system and possibly also not after 24 h in the stirred system. Fitting results are shown in Figure 4B and Table IB. The stirred system was fitted well by this model. For the e\e mixed system, C, (and therefore also Deff)could only be estimated with very high uncertainty (Table IB). Therefore, for the e\e mixed system the model was fitted again to the data with only the Deffas fitting parameter. C, was set to a value calculated from the KD obtained from fitting the stirred system. Now, estimated values of Deffdid not and 4.1 X m2/s for the stirred differ much: 6.4 X and the e\e mixed system, respectively. Compared to D,ff values reported in the literature, which were mainly in the range 10-13-10-16m2/s (27),our values are rather small. However, the values from the literature were obtained from desorption systems with equilibrium times of less than 2 days. According to the radial diffusion model, only 40% (e\e mixed) to 60% (stirred) of the equilibrium is achieved in the suspensions studied here within the experimental time. This suggests that the time to reach equilibrium is in the order of weeks to months. Equilibrium times of several months up to more than 1year have been reported for herbicides (39,40)and several polychlorinated biphenyl congeners (41). Therefore, our estimated values for C, and Deffdo not seem unreasonable. Calculated values of intraparticle porosity obtained from the estimated values of Deffand eq 3 were in the order of 0.003 (e\e mixed) and 0.004 (stirred), 1-2 orders smaller than values reported in the literature (27). Therefore, for the systems studied here, eq 3 and calculated porosity values appear to have no physical meaning, indicating that other retardation mechanisms also were affecting desorption kinetics. The nature of these mechanisms, such as for example hindered diffusion in micropores (22),remains unclear. 1352
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....
.....________
@--.-....
0
1
2
3
4
5 6 time (days)
~
7
8
9
10
0
Figure 5 . Desorption and biodegradation of a-, p- and y H C H as a function of time in a stirred contaminated soil suspension. After day 2, biodegradation rates were limited by desorption rates.
An F test was applied to compare the two desorption models. For the stirred system, it was found that both models were not significantly different in describing HCH desorption kinetics. However, for the e\e mixed system, the radial diffusion model fitted the data significantly better ( p = 0.05). Since this model accounts for an intraparticle mass-transfer resistance as a function of the aggregate radius, it can describe desorption in the two systems with only one parameter: Deff= (5.3 f 1.2) X mz/s. Though the radial diffusion model has the highest explanatory value, the first-order model remains useful because it is much easier to apply. For sorption of hydrophobic compounds, organic matter is often assumed to be the only effective sorbent (20,211. If KDis calculated from reported log K, values of 3.30 (42), a value of 30 cm3/g is obtained (KD= fdW, f , = 0,011). This value is 3 times smaller then the values estimated by applying the radial diffusion model. This indicates that in our systems soil constituents other than organic matter, i.e., calcareous materials, contribute to the sorption of a-HCH. In fact, our original contaminated soil contained large aggregates (>2 cm) with more than 80% carbonate and over 3 times the average a-HCH concentration. 2. Desorption and Intraparticle Mass-Transfer Effects on Biodegradation. Desorption of HCH together with its concurrent biodegradation were measured with time in an e\e mixed nonsterile soil slurry at 21 "C (Figure 5). For all three HCH isomers present in the contaminated soil, a fast initial increase in dissolved HCH was measured in the bulk. From the 240 mg/kg a-HCH originally present in the soil, - 4 mg/kg desorbed within the first 5 h of the experiment. This fast initial increase in dissolved HCH was followed by fast disappearance of aand y-HCH. The decrease of dissolved a- and r-HCH from solution was paralleled by HCH mineralization, as measured by chloride release. P-HCH proved to be recalcitrant as reported earlier ( 4 , 5 ) . After 2 days, a- and y-HCH concentrations in the bulk already dropped close to detection limit. This low solution concentration and the continued HCH mineralization, though a t lower rates (after 3 days), indicates that intraparticle or external mass transfer limited the overall bioconversion rates. Surfacechemical desorption reactions are not likely to be limiting at these slow bioconversion rates (28,29).Possible external mass-tranfer limitations were tested in the slurry system as follows: in end-over-end mixed slurries incubated at 30 "C, maximum a-HCH bioconversion rates (ro) were determined a t rotation rates of 5 and 10 rpm. Maximum bioconversion rates were found to be 38 f 6 and 45 f 6 mgkg-l-day-l, respectively. Assuming stationary diffusion through a nonturbulent liquid layer around the aggregates to be the external mass-transfer mechanism controlling
300
4
A
I
300
i 0
time (days1 Figure 6. Comparison of biodegradation rates of a-hexachlorocycb hexane in an end-over-end mixed contaminated soil suspension system (solid line) with that in a stirred contaminated soil suspension (dotted line) after different extents of a-HCH biomineralization. I
120 1
(x-HCH concentmtion (mg kgll
Figure 7. Comparison of maximum desorption (A) and biodegradation (6)rates ( T o ) of ahexachlorocyclohexane at different sorbed a-hexachktrocyclohexane concentrations for a stirred (S) and an end-overend mixed (E) soil suspension system.
bioconversion, a minimal nonturbulent layer thickness of 2 cm was calculated (AC = 1 mg/L, d, = 182 pm, D, = m2/s, ro = 40 mgmkg-l-day-'). This is an unrealistic value for mixed aqueous systems. Therefore, it was concluded that for the slurry systems used in this study external mass-transfer resistances can be neglected and that intraaggregate processes were controlling bioconversion. The slight increase of r, with mixing speed was probably the result of other factors, Le., decreasing aggregate sizes with increasing mixing rates. The limitation of biodegration by intraparticle mass transfer was shown by switching the e\e mixed suspension to a stirred system (Figure 6). Stirring the suspension resulted in the breakup of the aggregates, which led to a smaller average size of the agglomerates (Figure l), so reducing the average transport path length to the bulk liquid and exposing more adsorbed HCH to the bulk liquid. This resulted in enhanced desorption kinetics, as was already shown in the sterile desorption experiments (Figures 3 and 4), thus making higher biodegradation rates possible. 3. Desorption Kinetics versus Biodegradation Kinetics. Maximal (initial) rates of desorption of a-HCH in a sterile system were compared with its maximal rates of biodegradation at different a-HCH concentrations (Figure 7). At all a-HCH concentrations studied, desorption rates in the stirred suspension were consistently higher than in the e\e mixed system (Figure 7A). This shows that the effect of intraparticle mass transfer is also apparent at lower sorbed concentrations. As expected, initial rates of biodegradation (Figure 7B) were consistently higher than initial rates of desorption
-- -
rndial diffusion first order
k = 0.27 day.'
2
4
6
8 10 time (dnys)
12
14
16
18
Flgure 8. Observed and predicted (lines) biodegradation kinetics of a-hexachlorocyclohexane in end-over-end mixed (E) and stirred (S) soil suspensions. The models applied are a sorption-retarded radial diffusion model (solid lines) and a first-order model (dotted lines).
(Figure 7A) at all a-HCH concentrations studied except at the initial HCH concentration in the stirred system. This supports the concept that in the nonsterile systems a-HCH-degrading microorganisms were able to enhance desorption. Due to the microbial uptake of dissolved HCH, intraparticle mass-transfer rates were likely to become faster as a result of steeper concentration gradients. In addition, microorganisms might have been able to penetrate the aggregates to some extent. Thus, they can increase desorption rates by reducing mass-transport lengths. For one exceptional case, namely, at the highest initial sorbed concentration in the stirred system, the initial desorption rate was found to be higher than the maximum bioconversion rate (Figure 7A and B). This can be explained by the following. To obtain different initial sorbed concentrations, all suspensions were preincubated the same way as was done for the desorption experiments. The slurry with the highest initial sorbed concentration was not preincubated. Therefore, the easiest desorbable part of HCH, which was already degraded in all the preincubated soil suspensions, was still present in this suspension system. This also explains why the first part of HCH bioconversion was not limited by desorption (Figure 5): microbial activity was then still rather small while desorption rates were relatively high. 4. Biodegradation Kinetics in Terms of Intraparticle Mass Transfer. If intraparticle mass transfer in general governs the overall desorption kinetics and the potential biodegradation of a-HCH is much faster and therefore not limiting (Figure 5), it should be possible to describe the kinetics of biomineralization the same way as intraparticle mass transfer. Therefore, we tried to simulate the biokinetic data with the radial diffusion model and a first-order model. All data were obtained under conditions identical with those for the desorption study except that microbial activity was not inhibited. By applying the appropriate analytical solution of eq 2 (37) and the NLR procedure and by assuming that as a boundary condition the concentration at the location of the microbes is almost zero and that the volumetric-meanaggregate size corresponds with the diffusion domain with values for Defl obtained from fitting the desorption kinetics, the radial diffusion model could not be fitted to the kinetic data of bioconversion (not shown). A possible explanation of this "miss-fit" might be that the microbes were able to penetrate, at least partially, the aggregates. Therefore, instead of the fixed values of the volumetric-mean aggregate size, an average intraparticle diffusion distance (6) was taken as an independent model parameter. Figure 8 shows that the model now describes bioconversion kinetic data quite Environ. Sci. Technol., Vol. 24, No. 9, 1990
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well. Estimated values of 6 were 14 and 18 pm for the stirred and e\e mixed system, respectively. Application of a first-order model to the bioconversion kinetic data (by using the LRNI procedure) resulted in first-order rate constants of 0.42 and 0.27 day-' with r2 of 0.95 and 0.92 for the stirred and the e\e mixed systems, respectively. As can be seen in Figure 8, first-order and radial diffusion kinetics do not differ very much. Although radial mass-transfer mechanistics were most likely determining bioconversion kinetics, from an engineering point of the view the first-order model remains a useful tool in assessing the overall bioconversion phenomena for soil systems as complex as described in this study. Conclusions
A laboratory study was performed to investigate the effects of desorption processes on the kinetics of biomineralization of a-hexachlorocyclohexane in soil contaminated with toxic waste at a temperature of 20 OC. It was found that in mixed soil suspensions intraparticle masstransfer processes limited the kinetics of desorption and biodegradation. These results show that not only microbial activities but also desorption or mass-transfer processes can control the bioconversion of HCH and similar xenobiotic compounds in soil and sediment environments. Acknowledgments
We thank R. S. Summers for his thoughtful suggestions on the manuscript and T. N. P. Bosma for his assistance with the writing of the FORTRAN algorithms and the handling of the BMDP software. Registry No. a-HCH, 319-84-6; (3-HCH, 319-85-7; y-HCH, 58-89-9.
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Received for review May 16, 1989. Accepted April 27, 1990. This work was financially supported by a grant f r o m t h e Dutch Ministry of Housing, Physical Planning and Environment.
