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DIVISION S-2-SOIL CHEMISTRY

Adsorption Effects on Kinetics of Aldicarb Degradation

Equilibrium Model and Application to Incubation and Transport Experiments

^a Dep. of Soil, Crop and Atmospheric Sciences, Cornell Univ., Ithaca, NY 14853 USA
^b Dep. of Environmental Sciences, Univ. of California, Riverside, CA 92521 USA

leiguo{at}mail.ucr.edu

	ABSTRACT

TOP ABSTRACT INTRODUCTION Materials and methods Results and discussion Conclusions REFERENCES

 
The assumption of equilibrium adsorption was applied to both batch incubation and soil column leaching experiments to estimate degradation rate constants of the pesticide aldicarb [2-methyl-2-(methylthio) propionaldehyde o-(methylcarbamoyl) oxim] in the liquid (µ_l) and sorbed (µ_s) phases. The estimation was based on the observed K_d–µ relationship yielded in a soil amended with various amounts of activated C (AC), where K_d is the adsorption coefficient, and µ is the composed degradation rate constant from both phases. The inverse dependence of aldicarb degradation on K_d reveals that µ_l is faster than µ_s. For batch incubation experiments, the calculated µ_l and µ_s were 0.1228 and 0.0019 d^-1, respectively, differing by a factor of 65. In continuous-flow columns, µ_l and µ_s were both increased, with an estimated value of 0.2063 and 0.0055 d^-1, respectively, resulting in an accelerated overall degradation rate of aldicarb by 181% compared with the batch reactors. The results of our study indicate that, although degradation of aldicarb occurred primarily in the soil solution, it did not cease completely on the sorbed chemicals. The relative contributions of the two phases to the total degradation were therefore dependent on both the adsorption coefficient and the relative degradation rate constants for the dissolved and sorbed chemicals.

Abbreviations: AC, activated C • BTC, breakthrough curve • HPLC, high pressure liquid chromatography

	INTRODUCTION

TOP ABSTRACT INTRODUCTION Materials and methods Results and discussion Conclusions REFERENCES

 
SORPTION OF PESTICIDES in soil or other porous media is recognized to be an important process regulating pesticide transport and degradation in the environment (Jury et al., 1987; van Genuchten and Wagenet, 1989; Wagenet and Rao, 1990). The negative correlation between sorption and mobility has been well established (Lichtenstein, 1958; Helling, 1970; Hornsby and Davidson, 1973; Selim, 1989; Pignatello et al., 1993). However, the effect of sorption on degradation is much more complicated and depends on many factors related to microbial, soil, and environmental conditions and on the properties of the chemical of interest (Rao et al., 1993; Scow, 1993; Alexander, 1994). Completely opposite impacts have been observed for chemicals with different degradation routes and mechanisms. For example, Armstrong and Chesters (1969) reported that degradation of atrazine was accelerated by adsorption to clay minerals, while Ogram et al. (1985) found degradation of 2,4-D was inhibited by adsorption.

In general, sorption is often considered a process that limits pesticide degradation (Dao and Lavy, 1987; Guerin and Boyd, 1992; Ainsworth et al., 1993; Gamerdinger et al., 1997). This can be understood on the basis of the reduced bioavailability of sorbed compounds to soil microorganisms. However, in the field, sorption might increase degradation as a whole by increasing the residence time of pesticides in the root zone where most microbial activity occurs.

The relationship between adsorption and degradation of chemicals obviously is a fundamental one underlying the environmental behavior of pesticides and other chemical contaminants. However, although there are numerous studies focusing on the effects of adsorption on chemical bioavailability (e.g., Smith et al., 1992; Siahpush et al., 1992; Novak et al., 1995; Fu et al., 1996), very few studies have reported the respective degradation rates in soil solution and sorbed phases. Some studies intended to model chemical degradation under different sorption conditions automatically assumed that the degradation rate for sorbed compounds is zero (Mihelcic and Luthy, 1991; Fu et al., 1996; Shelton and Doherty, 1997). Scow et al. (1986) presented a two-compartment model, which explicitly included individual degradation rate constants for the sorbed and dissolved compounds. Unfortunately, the value of these rate constants could not be obtained from the experimental data due to the uncertainty associated with the inverse parameter estimation method adopted.

The purpose of our study was to evaluate quantitatively the effect of adsorption on degradation of the pesticide aldicarb. We present an equilibrium model that relates the degradation kinetics directly to adsorption coefficients. The model was applied to the degradation data of aldicarb obtained from both static incubation reactors and continuous-flow soil columns to estimate the respective degradation rates of the pesticide in the liquid and sorbed phases. In particular, the degradation rate constants calculated from these two sets of experiments were compared to evaluate the influence of water flow on degradation kinetics.

	Materials and methods

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Materials
The soil used in this study was a Valois silt loam (coarse-loamy, mixed, mesic Aeric Fragiaquept). It contained 16.4 g kg^-1 organic C, 301 g kg^-1 sand, 552 g kg^-1 silt, 147 g kg^-1 clay, and had a pH of 5.9. Prior to use, the soil was sieved through a 1-mm screen and was stored in polyethylene bags at room temperature. Powdered decolorizing AC was chosen as a soil amendment to modify the soil adsorption properties. Various amounts of AC were added to the soil, resulting in five amendment rates at 0 (AC0), 102 (AC1), 1025 (AC2), 2050 (AC3), and 5125 (AC4) µg g^-1, respectively.

Analytical standard aldicarb (purity 99.8%) was obtained from Rhône-Poulenc Ag Company (Research Triangle Park, NC). The pesticide has a water solubility of 6000 µg mL^-1 and is reported to have a low adsorption coefficient (K_d <1 mL g^-1; Montgomery, 1993). Aldicarb was selected because of its reported equilibrium transport behavior in a similar soil (Zhong et al., 1986) and its relatively short persistence.

Batch Sorption Experiments
Measurement of aldicarb sorption isotherms and kinetic sorption were carried out in triplicate at 24 ± 1°C. The experiments were performed using 20-mL scintillation vials containing soil–water suspensions. The suspensions were composed of 5 g soil and 10 mL 0.005 M CaSO₄ containing 200 µg mL^-1 of HgCl₂ as biocide. The sorption isotherms were measured using six initial aldicarb concentrations: 5, 10, 20, 40, 60, and 100 µg mL^-1 and a shaking period of 72 to 240 h, depending on results of the kinetic experiments.

The kinetic experiments were conducted only with 20 µg mL^-1 aldicarb solution. Conditions for the kinetic experiments were identical to the isotherm experiments, except for the shaking time, which lasted 24, 48, 72, 100, or 240 h. Shaking was continuous for up to 48 h and then 12 h d^-1 for longer shaking periods.

At the end of shaking, the soil–water suspension was allowed to settle for 10 to 15 min. A 2-mL aliquot of the supernatant was then filtered through a 0.45-µm Nylon-66 membrane filter with polypropylene housing (Fisher Scientific, Pittsburgh, PA). The aldicarb concentration of the filtrate was determined by high pressure liquid chromatography (HPLC). The sorbed concentration was calculated as the difference between the initial concentration and the concentration at equilibrium. Losses to vials and during filtration were determined to be negligible from the blank samples that contained only the pesticide solution.

Incubation Experiments
Two sets of batch incubation experiments were conducted to measure the total soil respiration and aldicarb degradation rate, respectively. The soil respiration experiments were carried out in duplicate in 250-mL sidearm biometers (Bellco Glass, Vineland, NJ) following a modified procedure as described by Bartha and Pramer (1965). These experiments measured total CO₂ evolved from the soil, which can be used to indicate overall microbial activity during the incubation period. The purpose of these experiments was to test whether the AC amendment supports microbial growth and thus changes the microbial activity.

Thirty grams of each soil amended with AC was added to the biometer. A 20-mL scintillation vial containing 5 mL of 0.5 M KOH was placed in the sidearm, and the biometer was immediately capped with rubber stops. The CO₂ evolved from the incubated soil sample was trapped by the KOH solution and the amount of CO₂ was determined by titration once every week. At each sampling, the scintillation vial was removed, and an aliquot (1 mL) of the trapping solution was titrated by 0.1 M HCl to quantify the absorbance of CO₂. The endpoint of titration was indicated by 1% (by weight) phenolphthalein.

The incubation experiments for aldicarb degradation were performed in triplicate in 250-mL flasks containing 200 g (dry basis) soil at saturated water content (38.8%). The soil was preconditioned for 4 wk prior to pesticide addition by adding an appropriate amount of distilled water, determined by weighing, to bring the water content to approximately the field capacity (28.8%). Two and one-half milliliters of the aldicarb stock solution (4042 µg mL^-1) was then added dropwise evenly across the surface of the soil, yielding an application rate of 50 µg g^-1. The soil was mixed carefully with a stainless steel spoon. The water content was then brought to saturation. The flask was left open for the first few days to allow excessive water to evaporate and was then covered with aluminum foil for incubation at the room temperature of 24 ± 2°C. Distilled water was added as needed during the incubation, and soil samples were removed periodically for analysis.

At each sampling event, 10 to 20 g of soil sample was transferred into a 60-mL high density polyethylene bottle, and was extracted following a modified procedure provided by the manufacturer (Rhône-Poulenc Ag Company). The soil sample was first shaken with 30 mL of methanol on a reciprocating shaker for 5 h. An aliquot (10 mL) of the extract was then filtered through a 0.45-µm nylon filter. The filtrate was transferred to a 15-mL graduated glass centrifuge tube, and was further condensed as needed under N₂. The aldicarb concentration of the condensed filtrate was determined by HPLC. Due to variation in the weight of soil samples, the water/methanol ratio of the extractant was calculated to be 0.21 to 0.30. Accordingly, the mass recovery of the extraction procedure was measured separately at two water/methanol ratios, 0.25 and 0.40, and was found to be indistinguishable for all five AC soil treatments (1.01 ± 0.028).

Soil Column Leaching Experiments
The soil column leaching experiments were designed to measure pesticide degradation during water flow. For this purpose, a long soil column (26 cm), combined with a low flow velocity (14 cm d^-1), was used to increase the residence time of aldicarb in the column. The glass column had an internal diameter of 5 cm, and was connected to a digitized isocratic LC pump (Perkin-Elmer Model 250, Perkin-Elmer Corp., Norwalk, CT) after being packed with each soil at a density of 1.30 g cm^-3. At this density, the measured porosity was 43%.

Each column was saturated with 0.005 M CaSO₄ from below at a very slow flow velocity. The column was then drained for 2 wk before saturation was resumed. After the flow rate was adjusted to a steady state, a pulse of one pore volume of pesticide solution containing 28 µg mL^-1 aldicarb was applied. The pesticide was subsequently displaced by pesticide-free CaSO₄ solution. The effluent discharged from the column was collected in distinct fractions and analyzed for aldicarb by HPLC. To facilitate aldicarb degradation through oxidation, the influent reservoir (12-L capacity) was connected to a constant air-flow source in order to maintain a high O₂ level in the influent solution. The estimated O₂ concentration was 315 µmole L^-1 and was calculated to be sufficient to oxidize twice the total aldicarb applied in the column to its primary oxidized product. At the termination of leaching, soil in the column was extruded by applying an air stream from one end. Soil samples were homogenized and an aliquot of 50 g was analyzed for the uneluted pesticide using a similar procedure as described for the samples of incubation experiments.

High Pressure Liquid Chromatography Procedures
The HPLC system was equipped with a Perkin Elmer LC Pump Series 410, an LC-95 UV/Visible Spectrophotometer Detector (Perkin Elmer, Norwalk, CT), a Hewlett Packard ODS Hypersil column (5 mm packing and 125 x 4 mm; Hewlett Packard Analytical, Palo Alto, CA) and a Rheodyne Model 7125 Syringe Injector with a 100-µl sample loop (Rheodyne, Cotati, CA). The eluting solvent was methanol-water (35:75) containing 4% (v/v) acetic acid. The flow rate of the eluting solvent was set at 1 mL min^-1. The retention time for aldicarb under these conditions was 8 min, and the detection limit was 0.2 µg mL^-1. The analytical precision of the procedure was ± 2.37% based on standard samples.

	Results and discussion

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Sorption Kinetics and Isotherms
Sorption of aldicarb at various times is shown in Fig. 1 . Sorption virtually reached equilibrium within the first 72 h in all soil treatments, except for AC4, in which sorption did not appear to reach equilibrium at 240 h. The sorption isotherms shown in Fig. 2 are the values measured at the shaking time of 72 h, with the exception of the treatment AC4, for which the sorption isotherm was measured at 240 h.

View larger version (13K): [in this window] [in a new window]   Fig. 1 Kinetics of aldicarb adsorption measured in a Valois silt loam amended with various amounts of activated C (error bars represent one standard deviation)

View larger version (18K): [in this window] [in a new window]   Fig. 2 Adsorption isotherm of aldicarb measured in a Valois silt loam amended with various amounts of activated carbon. Symbols denote measurements (error bars represent one standard deviation), and lines represent fitted linear Freundlich isotherms

The purpose of using organic amendment was to modify the sorption properties of the soil. We chose AC as the C source because sorption of organic chemicals is known to be very sensitive to AC. The aldicarb K_d increased from 0.36 to 47.4 mL g^-1 upon addition of 5000 µg g^-1 AC (Table 1). These adsorption coefficients represent 48 to 99% of total aldicarb in the sorbed form.

Soil Respiration
Total soil respiration, as measured by the amount of CO₂ evolved from the soil, is plotted in Fig. 3 . The data was calculated assuming that the absorption efficiency of CO₂ by KOH was 100%. The CO₂ evolution rate, varying between 0.97 to 4.81 µg C g^-1d^-1, was erratic, but overall rather stable for all soil treatments during the incubation period of 100 d. Based on an F test, there were no significant differences at P = 0.05 among various AC-amended soil treatments, which demonstrates that the activated C added could not support the microbial growth in the soil. This finding is important to the later analysis and conclusions of our study, because the relationship observed between adsorption and degradation would be attributed to totally different causes if microbial activity were changed in the presence of AC.

View larger version (13K): [in this window] [in a new window]   Fig. 3 Total soil respiration measured in a Valois silt loam amended with various amounts of activated carbon (error bars represent one standard deviation)

View larger version (22K): [in this window] [in a new window]   Fig. 4 Decay of aldicarb in a Valois silt loam amended with various amounts of activated carbon. Symbols denote measurements (error bars represent one standard deviation), and lines are fitted first-order reactions

View larger version (13K): [in this window] [in a new window]   Fig. 5 The Kd–µ relationship of aldicarb for incubation experiments

(1)

where C is the resident concentration of aldicarb, s is the sorbed concentration, c is the dissolved concentration, is the soil bulk density, and is the volumetric water content. Recognizing that for equilibrium adsorption

(2)

integration of Eq. [1] results in a composed first-order kinetic term describing overall aldicarb decay for the entire system:

(3)

where C₀ is the concentration at time zero, and µ is the composed degradation rate constant and is equal to:

(4)

Equation [3] suggests that if degradation in each phase follows first-order kinetics, the overall decay will also be a first-order reaction. Further characterization of µ_l and µ_s can be achieved by nonlinear regression of the observed µ–K_d relationship curve to Eq. [4]. The curve fitting results are also presented in Fig. 5. The estimated µ_l was 0.1228 d^-1, which was 65 times that calculated for µ_s (0.0019 d^-1). These results are consistent with the perception that adsorption of a chemical by soil reduces its availability to microbial uptake and thus reduces its degradation by microorganisms (Ainsworth et al., 1993; Scow, 1993).

A fundamental assumption underlying our analysis and conclusions is equilibrium adsorption. Apparently this assumption is not valid for the soil–solute systems examined in this study, especially in the presence of AC. Our kinetic sorption experiments (Fig. 1) indicated that adsorption of aldicarb required at least 24 h to reach a quasi-equilibrium state, and this time increases with AC rates. So the estimated µ_l and µ_s would have an error term which is directly related to the fraction of nonequilibrium adsorption in the total adsorption. The estimation of µ_l and µ_s under the assumption of nonequilibrium adsorption is addressed in a forthcoming paper (Guo et al., 1999). Nevertheless, the error is considered to be insignificant considering the time scale during which degradation was assessed and the relatively small fraction of kinetic adsorption at longer times.

Degradation in Continuous Flowing Soil Columns
The degradation rate constant of aldicarb in the continuous-flow soil columns was calculated using the following equation that we developed previously (Guo and Wagenet, 1999):

(5)

where M_r is the residual mass of aldicarb remained in soil after leaching, M₀ is the total mass applied into the soil column, f(t) is the travel time probability density function of aldicarb in the soil column, is the time equal to one-half pulse length (1/2T₀), and T is the experimental duration. With all the parameters determined experimentally, the degradation rate constant µ can be calculated by solving Eq. [5] numerically based on the measured breakthrough curve (BTC), which represents f(t).

The measured BTCs of aldicarb, along with the experimental parameters and the estimated µ, are presented in Fig. 6 for all soil columns. No BTC was available from the column of AC4, because no detectable aldicarb was found in the effluent samples despite that the length of the column was reduced to 5 cm. A simple calculation using a K_d of 47 (that measured for AC4 in the batch sorption experiments) indicates that about 35 pore volumes of water are needed for aldicarb to break through the column, assuming no degradation.

View larger version (24K): [in this window] [in a new window]   Fig. 6 Measured breakthrough curves of aldicarb. The residual mass (Mr) for AC0 and AC1 was below method detection limit (<0.04 µg g-1). L is the length of the column, v is the pore water velocity, and T0 is the step pulse length in the unit of pore volume

The measured BTCs of aldicarb did not show appreciably elevated tailing (Fig. 6), which suggests that equilibrium sorption prevailed during its transport through the columns. Results of the kinetic sorption experiments indicate that it took 72 h for aldicarb to reach sorption equilibrium in most of the soil treatments. Therefore, sorption achieved during this length of time is still appropriate to be viewed as "instantaneous" with respect to transport at the flow rate used in this study.

The µ–K_d curve obtained for the columns is shown in Fig. 7 . The relationship that was observed for incubation experiments still held under transport conditions: a higher K_d was associated with a slower degradation. This demonstrates again that the sorbed aldicarb was degraded more slowly than the dissolved aldicarb. The curve fitting analysis using Eq. [4] reveals that the accelerated degradation of aldicarb under transport conditions was attributable to both enhanced µ_l ( in the columns vs. 0.1228 d^-1 in the incubation flasks) and µ_s (0.0055 vs. 0.0019 d^-1). However, the increase in µ_s was larger than in µ_l. The difference between them reduced from 65-fold in static incubation experiments to 38-fold under transport conditions.

View larger version (12K): [in this window] [in a new window]   Fig. 7 The Kd–µ relationship of aldicarb in continuous-flow columns

	Conclusions

TOP ABSTRACT INTRODUCTION Materials and methods Results and discussion Conclusions REFERENCES

Due to the difference in the surface properties of the AC that we used in this study and soil organic matter or minerals, it is possible that the degradation rate constants we estimated in this study may deviate somewhat from those in natural soil. The degradation rates for a particular soil would depend on both the nature of the surfaces (most notably hydrophobic vs. hydrophilic surfaces) and the properties of the chemical under consideration.

	ACKNOWLEDGMENTS

 
This research was supported by funding from the U.S. Department of Agriculture, BARD Project no. US2268-93C.

Received for publication August 5, 1998.

	REFERENCES

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