Soil Science Society of America Journal 63:1077-1086 (1999)
© 1999 Soil Science Society of America
DIVISION S-1-SOIL PHYSICS
Transport of Triticonazole in Homogeneous Soil Columns
Influence of Nonequilibrium Sorption
C. Beigela and
L. Di Pietroa
a Unité de Science du Sol, I.N.R.A., Domaine St Paul, Site Agroparc, 84914 Avignon Cedex 9, France
lili{at}avignon.inra.fr
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ABSTRACT
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Nonequilibrium sorption of pesticides is frequently reported to greatly affect their transport and dissipation in soil. This study was aimed at evaluating the performances of equilibrium and two site–two region nonequilibrium convective–dispersive models for describing the sorption and decay characteristics during transport of triticonazole systemic fungicide in water-saturated homogeneous soil. Chloride and 14C-triticonazole column displacement experiments were carried out in a loamy clay soil under steady-state water flow at high pore water velocities. The symmetrical breakthrough curves (BTC) obtained with the conservative tracer showed no significant physical nonequilibrium and were used to estimate a dispersivity of 0.06 cm. Compared with chloride, the 14C-triticonazole BTC was strongly asymmetrical and shifted to the right, with a broad, extended tailing characteristic of sorption nonequilibrium. Chemical analysis of the soil after elution showed that bound residues were rapidly formed during transport. These bound residues were accounted for as decayed in the models. The two-site model correctly described the first part of the tailing, with an estimated partition coefficient Kd of 1.5 L kg-1 for instantaneous sorption, and it predicted high values in the range of 58 d-1, and 7 d-1 for the sorption and decay first-order rates, respectively. However, the model failed to describe the slower, extended release of 14C-triticonazole. Nonequilibrium sorption and formation of bound residues of triticonazole were attributed to the rate-limiting diffusive process. It was thus concluded that use of a single first-order rate constant for description and prediction of both nonequilibrium sorption and dissipation of triticonazole in soil is not appropriate.
Abbreviations: BTC, breakthrough curves • CDE, convective–dispersive equation • CDeq, CXTFIT 2.0 deterministic equilibrium CDE model • CDnoneq, CXTFIT 2.0 deterministic two-site, two-region nonequilibrium model • HOC, hydrophobic organic compounds • LEA, local equilibrium for solute adsorption • LSC, liquid scintillation counting • OM, organic matter • STD, standard deviation
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INTRODUCTION
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TRITICONAZOLE [(1RS)-(E)-5-(4-chlorophenyl-methylen)-2,2-dimethyl-1- (1H-1,2,4-triazol-1-ylmethyl)- cyclopentan-1-ol] is a new triazole systemic fungicide, developed by Rhône-Poulenc Agro (Courbevoie, France), that is used in cereal seed treatments. Triticonazole controls major seedborne, foliar, and straw diseases, thus allowing for cereal protection from seed to developed growth stages. The efficacy of systemic pesticides applied in seed treatments depends closely on their dissipation, localization, and availability in the soil profile in relation to their uptake by the plant root system. The transport and fate of these molecules in soils is, therefore, of crucial interest for an optimal utilization.
Solute transport of organic chemicals in soil depends on the soil's structural and hydraulic properties, and it is controlled by sorption and degradation, which both limit the mobility of the pesticides in soil. These major dissipation processes have been extensively studied (Graham-Bryce, 1981; Weber and Miller, 1989; Weber, 1991). The difficulty of accurately accounting for these two interacting processes with simple input parameters is one of the main problems encountered in predicting and modeling solute transport in soil (Calvet, 1995).
Sorption of hydrophobic organic compounds (HOC) has been related to soil organic matter (OM) through nonspecific interaction mechanisms (Hamaker and Thompson, 1972; Ainsworth et al., 1989; Barriuso and Calvet, 1992). Soil sorption isotherms of HOC at diluted concentrations are usually linear (Calvet, 1989), and a soil–solution partition coefficient Kd (in L kg-1 for sorbed concentration Se divided by solution concentration Ce at equilibrium) is frequently used to characterize HOC sorption on a particular soil. These coefficients are experimentally determined in batch systems, assuming reversible, quasi-instantaneous sorption. On the other hand, there is increasing evidence of nonreversible and time-dependent sorption of most pesticides in soil (Lehmann et al., 1990; Scribner et al., 1992, Barriuso et al., 1992). Sorption kinetics for HOC usually exhibit a two-stage approach to equilibrium, with a short initial phase of rapid uptake followed by an extended period of much slower uptake (Lee et al., 1988; Gaston and Locke, 1995).
The degradation of pesticides in soil is mainly due to biological transformations and is thus controlled by the availability of the organic chemical and by the activity of the soil microflora (Torstensson, 1987). It is usually characterized by first-order rate constants (k) calculated from mineralization or dissipation data, using the first-order relation
are the substrate concentrations in soil at time 0 and t, respectively. However, degradation may not always follow simple first-order kinetics if the pesticide is subject to significant abiotic degradation. Also, the degrading capacity of the soil microflora may vary with time, because the growth and activity of the degrading microorganisms are extremely sensitive to environmental conditions such as temperature and humidity (Walker et al. 1992, Veeh et al. 1996), and because adaptation phenomena may occur (Felsot and Shelton, 1993; Ou et al., 1993). Furthermore, changes in the pesticide availability with time that are due to nonequilibrium sorption may also affect the degradation rate.
The sorption and degradation characteristics of triticonazole systemic fungicide in a loamy soil of Grignon, France, (Typic Eutrochrept) have been examined in previous studies (Beigel et al., 1997, 1999). The degradation of triticonazole was essentially due to microbial, cometabolic transformations that could be adequately characterized by first-order mineralization rate constants ranging from 0.3 x 10-3 to 0.6 x 10-3 d-1, depending on the initial dose applied. Batch equilibrium studies showed that triticonazole equilibrium sorption was related to soil organic matter content and could be approximated by a linear isotherm, with a measured Kd of 4.35 L kg-1; however, it appeared that triticonazole sorption into Grignon soil during incubation and diffusion experiments was strongly time-dependent and increased as the time of contact with the soil increased. A time-dependent, apparent partition coefficient was measured, which increased from 2.5 to 10 L kg-1 during a 130-d incubation. Rate-limiting desorption, and formation of methanol nonextractable, bound residues were observed, which resulted in a decreasing apparent diffusion rate of triticonazole in soil. This was attributed to rate-limiting intrasorbent (organic matter) diffusion to restricted sorption sites. Use of the batch equilibrium Kd to characterize triticonazole sorption and desorption during transport would thus prove erroneous, and nonequilibrium sorption conditions need to be accounted for.
The objective of the present study was to test simple conceptual coupled-process models to describe the one-dimensional transport of triticonazole in homogeneous saturated soil columns. We evaluated the deterministic equilibrium convective–dispersive equation (CDE) and the two-site, two-region, deterministic, nonequilibrium CDE models using the computer program CXTFIT 2.0 of Toride et al. (1995), which is an updated version of the CXTFIT code of Parker and van Genuchten (1984a).
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Materials and methods
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Soil and Chemicals
The soil used in this study was a loamy clay (Typic Eutrochrept), sampled in the surface layer (0–20 cm) of a continuous wheat experimental plot located at Grignon, France. It had a pH in water of 8.2, with the composition: 29.1% clay, 54.0% silt, 14.5% sand, and 1.04% organic C. Soil samples were air dried and passed through a 2-mm sieve. To avoid problems of clogging while packing the columns and during elution, the finer particles were removed by further sieving at 0.5-mm. Soil water content (kg kg-1) was 0.043%.
Carbon-14-U-benzyl-labeled triticonazole (specific activity: 1184 MBq mmol-1; radiopurity >98%) was provided by Rhône-Poulenc Agrochemicals, Lyon, France. Triticonazole water solubility is 8.4 mg L-1 and 10.6 mg L-1 at 20 and 22°C, respectively, and the distribution coefficient between octanol and water is 1950. A solution of 14C-triticonazole at 5.0 mg L-1 and 8.332 Kbq mL-1 (0.224 µCi mL-1) was prepared for input tracer solution by adding 14C-triticonazole methanol stock solution to a saturated water solution of triticonazole (analytical standard, purity >92%) and adjusting the concentration with MilliQ water (Millipore, Saint-Quentin, France). The solution concentrations were measured at 262.5 nm with a UV-visible spectrophotometer Lambda V (Perkin-Elmer, Überlingen, Germany).1
Chloride was used as a nonreactive, conservative tracer to determine the soil hydrodynamic dispersion coefficient at different water velocities. A solution of CaCl2 with Cl- at 1 g L-1 was prepared for input tracer solution by diluting CaCl2,2H2O (analytical reagent >98% purity; R.P. Normapur, Prolabo, Paris, France) in the proper amount of MilliQ water.
Miscible Displacement Experiments
The BTC for CaCl2 and triticonazole were measured in water saturated, isotropic homogeneous soil columns. The system consisted of PVC columns 65 mm in length (L) and 55 mm diam. A stainless steel porous filter of 50-µm mesh was used as bed support on the bottom of the column. At the top of the column, a void volume 5 mm deep was provided as a mixing cell to allow for the formation of a piston-like water front. The columns were packed under water saturation conditions in a water bath by adding successive layers of soil to establish uniform bulk density and water content. The columns were then covered to avoid evaporation and allowed to equilibrate in the water bath for 24 h at 22 ± 2°C. Measured in 10 test columns, mean gravimetric water content (
g), volumetric water content (v), and bulk density (Mg m-3) were 0.65 ± 0.01, 0.60 ± 0.01, and 1.07 ± 0.01, respectively. The pore volume (V0) was calculated as the product of the column volume and v at 92.7 cm3. Soil saturated hydraulic conductivity (Ksat), measured in similar soil conditions, was of 97.2 cm d-1.
A constant pressure head (
h) was applied on the soil columns for steady-state flow conditions, using the system shown in Fig. 1
. The desired pore water velocities were obtained by varying the level of the column relative to the overflow outlet. The elution fluxes at the columns outlet (Q, cm3 d-1) were measured during the elution experiments by weighing the amount of solution eluted at regular time intervals. The Darcy's velocities (q, cm d-1) were evaluated by dividing Q by the cross sectional area of the soil column (23.76 cm2). We observed a good agreement between the measured and theoretical calculated Darcy's velocities for different h. All the elution experiments were performed in duplicate at 22 ± 2 °C. For the chloride-miscible displacement studies, three different q of 125, 160, and 212.5 cm d-1, corresponding to h of 2, 4, and 8 cm were established, respectively. The average pore water velocities (v) for the duplicates were calculated from
at 208, 263, and 354 cm d-1, respectively. The elution of triticonazole was performed at the lowest chloride Darcian velocity of 125 cm d-1. The columns were first supplied with MilliQ water until steady-state flow conditions were attained at the desired infiltration rate. A pulse of the tracer solution corresponding to 0.5 pore volume was then applied at the same rate, and the sytem was then switched back to the water reservoir for elution periods of 3 h (chloride BTC), and 11 h (triticonazole BTC). At regular time intervals, aliquot samples of the leachate were collected, weighed, and stored at -20°C until chemical analysis. At the end of the elution time course, sequential sampling of the triticonazole soil columns was performed by extruding and slicing the soil in 10 incremental disks of 6-mm sections using a soil extruding screw procedure described in a previous paper (Beigel et al. 1997).
Chemical Analysis
Chloride concentration in the effluent samples was measured by capillary analysis by direct UV detection with a sulfate–OFM–CHES electrolyte, using a CIA analyzer equipped with a AccuSep 75 µm x 60 cm capillary and Millenium 2.1 software (Waters, Milford, MA).
Triticonazole concentration in the effluent samples was measured by liquid scintillation counting (LSC). Aliquots of the elution samples (0.5 mL) were pipetted and put in scintillation vials. Four milliliters of Ultima Gold XR LSC Cocktail (Packard Instrument, Meriden, CT) were added, and the amount of radioactivity was measured by LSC using a Kontron Betamatic V counter (Kontron Instrument, Montigny le Bretonneux, France).
Total 14C-triticonazole residues remaining in the soil samples were measured by LSC of the 14CO2 evolved after combustion of triplicate 300-mg aliquots of air-dried and finely ground soil with a Sample Oxidizer 307 (Packard Instrument, Meriden, CT). Extractable residues were analyzed after exhaustive extraction with methanol. The soil samples were extracted twice with 50 mL of methanol. After 24 h of shaking, the samples were centrifuged for 15 min at 5000 rpm (8000 g), and the radioactivity content in the supernatant was measured by LSC as previously described. After the second extraction, the soil pellets were air dried, and the remaining radioactivity in the soil was measured by combustion of triplicate 300-mg soil aliquots as previously described.
Models
CXTFIT 2.0 (Torride et al., 1995) is a program presenting a number of analytical solutions for one-dimensional transport models based on the convection–dispersion equation (CDE). Assuming steady-state flow in a homogeneous soil and first-order transformation kinetics with uniformly distributed nongrowing biomass, the equilibrium CDE may be written as
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where t is time (d), x is depth (cm), 
is soil bulk density (g cm-3), is soil volumetric water content (cm3 cm-3), C is the concentration of the liquid phase (mg L-1), S is the concentration of the adsorbed phase (mg kg-1), v is the average pore water velocity (cm d-1), D is the hydrodynamic dispersion coefficient (cm2 d-1), and µliq is a first-order decay coefficient for degradation in the liquid phase (d-1). Depending on the equilibrium–nonequilibrium assumptions about S, we used CXTFIT 2.0 deterministic equilibrium CDE model (Mode 1), further noted as CDeq, or CXTFIT 2.0 deterministic two-site, two-region nonequilibrium model (Mode 2), further noted as CDnoneq.
The CDeq model assumes local equilibrium (LEA) for solute adsorption and that sorption can be described by a single linear isotherm,
, where Se and Ce are the concentrations in sorbed and liquid phases at equilibrium, and Kd is the equilibrium partition coefficient (L kg-1).
The two-site, two-region bicontinuum model has been formulated to account for either sorption-related or transport-related nonequilibrium during solute transport. The two-site nonequilibrium concept assumes that sorption sites in soils can be classified into two fractions. In the first fraction, sorption is instantaneous and is described by an equilibrium sorption isotherm (Type 1, equilibrium). In the second fraction, sorption is time-dependent and follows first-order kinetics (Type 2, kinetic). In this case, the rate-limiting step for Type 2 sites would be either chemical (chemisorption), or diffusive (intraparticle or intrasorbent diffusive mass transfer), as discussed by Brusseau et al. (1991). The two-region approach assumes that the liquid phase can be partitioned into mobile (flowing, macropore domain) and immobile (stagnant, matrix, or micropore domain) regions. The exchange between the two liquid regions is modeled by a first-order kinetic equation. Flow occurs only in the mobile region. Sorption is assumed to be instantaneous on all sorption sites, and the sorption rate is limited here by the diffusion of the solutes to the exchange sites in the stagnant phase. If dimensionless parameters are employed, then the two-site and two-region models reduce to the same dimensionless form:
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 | (3) |
where
, P is the Peclet number
, R is the retardation factor, defined as
. The subscripts 1 and 2 refer to equilibrium and nonequilibrium sites respectively, ß is a fraction factor, and 
is a dimensionless, mass transfer coefficient. The various dimensionless parameters have different meanings for the two-site and two-region models, which are defined in the CXTFIT 2.0 code (Toride et al., 1995).
For the two-site model, ß and are defined as
 | (4) |
 | (5) |
where f is the fraction of Type 1 sites, and 
(d-1) is the first-order rate for the kinetic, Type 2 sites. The dimensionless decay terms µ1, for degradation in the liquid phase only, reduces to
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For the two-region model, ß and are defined as
 | (7) |
where m is the volumetric water content of the mobile region, and f is the fraction of sorption sites in the mobile region; and
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where is a first-order mass coefficient governing the rate of solute exchange between the mobile and immobile liquid regions.
Parameters Estimation
The hydrodynamic dispersion coefficient D of the soil was estimated from the chloride data with both models using the nonlinear least-squares parameter optimization method (inverse problem). The solute transport parameters v and D at the three infiltration rates were first estimated from the chloride BTC (pooled results from the duplicate columns) with the CDeq model with R set to 1. The two-region approach from the CDnoneq was then used to check for potential physical nonequilibrium. Nonequilibrium transport parameters D, , and ß were estimated for chloride nonreactive tracer, with R set to 1, and v set to the values obtained from CDeq. The values of D obtained from the chloride BTC were used to estimate the soil dispersivity,
. The dispersivity was used to calculate the value of D for triticonazole BTC, which was then introduced in CDeq to estimate the retardation factor R, the dimensionless parameter for linear sorption, and the first-order degradation rate µliq from the triticonazole BTC. CDnoneq was used to estimate the retardation factor, and the nonequilibrium dimensionless parameters ß, , and µ for triticonazole transport and decay.
Both models were used under the conditions of semi-infinite system, first-type (flux) boundary conditions, pulse input, linear sorption and first-order degradation in the liquid phase only. All these conditions were assumed to reasonably apply to our elution column experimental setup (see Parker and van Genuchten, 1984b) and to triticonazole sorption and degradation characteristics. The corresponding initial and boundary conditions, as well as some analytical solutions are detailed in the CXTFIT 2.0 code (Toride et al., 1995).
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Results
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Chloride Breakthrough Curves
The BTC of chloride conservative tracer measured at the three water velocities were almost identical (Fig. 2)
. The effluent curves appeared symmetrical and sigmoidal, with invariant frontal and distal portions. No significant retardation occured, as chloride was detected in the effluent after application of 0.8 V/V0 of water, and relative concentration peaks were measured at 1.25 V/V0.
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Fig. 2 Effect of pore water velocity on the measured (symbols) and simulated (lines) breakthrough curves in Grignon soil for displacement through Grignon soil of a pulse of 0.5 V0 of chloride conservative tracer at 1 g L-1
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The curve fitting results of CDeq and of the two-region CDnoneq to the experimental chloride data are listed in Table 1
. The CDE fitted the observed results well, as indicated by the high correlation coefficient r values obtained (>0.97). Estimates of v with CDeq for the duplicate columns at the three infiltration rates were close to the measured velocities; however, the estimation of the dispersion coefficients D was not satisfactory, as the standard deviation (STD) values were higher than the D values. Use of CDnoneq with the CDeq fitted values for v proved more efficient for the estimation of D, as STD were much lower. Estimated values for D were very low and increased from 16 to 25 cm2 d-1 with increasing water velocity. As D appeared directly proportional to v, the dimensionless Peclet number
, and the dispersivity
were almost invariant, with mean values of 86.7 ± 5.5 for P, and 0.06 ± 0.02 cm for 
, respectively. The estimated parameter ß for physical nonequilibrium appeared dependent upon the water velocity, with ß decreasing from 1 to 0.92 with estimated v values decreasing from 355 to 217 cm d-1.
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Table 1 Comparison of CDeq and CDnoneq parameter values estimated from the chloride data at three pore water velocities
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Triticonazole Breakthrough Curves
Comparison of the breakthrough of chloride and triticonazole measured at a water velocity of 208 cm d-1 for a pulse of 0.5 V0 is shown in Fig. 3
. Triticonazole BTC appeared shifted to the right. The relative concentration of triticonazole in the effluent was comparatively much lower than the conservative tracer concentration. A delayed concentration peak at a low C/C0 of 0.16 was obtained after elution of 
3 pore volumes. Triticonazole BTC was strongly asymetrical in shape (Fig. 4)
. The distal part of the elution peak appeared skewed and biphasic, showing a rapid release in the early tailing followed by a much slower release in the extended tailing.
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Fig. 3 Comparison of chloride (line) and 14C-triticonazole (symbols) measured breakthrough curves for displacement through Grignon soil of a pulse of 0.5 V0 and a pore water velocity of 208 cm d-1. Initial concentrations of the tracers input solutions were 1 g L-1 for chloride and 5 mg L-1 for triticonazole
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Fig. 4 Measured and CDeq-simulated breakthrough curves for displacement through Grignon soil of a pulse of 0.5 V0 of 14C-triticonazole at 5 mg L-1. Simulated curves were calculated with the model parameters v and D set to 208 cm d-1 and 12.48 cm2 d-1, respectively, without decay (dotted line) or with first-order decay in the liquid phase (solid line)
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Results of CDeq and CDnoneq fits with v set to 208 cm d-1, and
set to 12.48 cm2 d-1 are summarized in Table 2
. The equilibrium model proved inefficient in describing the experimental data (Fig. 4). CDeq could not account for any of the asymmetry and tailing, and it could only describe the position of the delayed concentration peak; i.e. the retardation due to instantaneous equilibrium sorption. The CDeq fit without decay resulted in a peak twice as high as the observed data. The introduction of a first-order decay constant (sink term) improved the fit, as it allowed better description of the spreading of the elution peak. Use of the two-site nonequilibrium model considerably improved the fit to the experimental BTC (r > 0.93). CDnoneq accounted for the early tailing asymmetry of the concentration peak, but it still failed to account for the extended tailing of the BTC (Fig. 5)
. The model predicted that all the applied radioactivity would be recovered in the effluent after elution of 12 pore volumes of water, whereas significant amounts of 14C were still detected in the effluent after elution of the 16 pore volumes. CDnoneq without decay slightly overpredicted the width of the concentration peak. Introduction of the decay term improved the fit (r of 0.97), as it allowed better description of the spreading of the elution peak. The dimensionless parameter for decay µ1 allows for the calculation of the first-order degradation rate in the soil liquid phase, µliq. The estimated value of µliq for the degradation of triticonazole was 7 d-1.
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Table 2 Comparison of CDeq and CDnoneq parameter values estimated from the 14C-triticonazole data, with v and D set to 208 cm d-1 and 12.48 cm2 d-1, respectively
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Fig. 5 Measured and CDnoneq-simulated breakthrough curves for displacement through Grignon soil of a pulse of 0.5 V0 of 14C-triticonazole at 5 mg L-1. Simulated curves were calculated with the model parameters v and D set to 208 cm d-1 and 12.48 cm2 d-1, respectively, without decay (dotted line) or with first-order decay in the liquid phase (solid line)
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In both models, instantaneous sorption is characterized by the dimensionless retardation factor R, based on the mean position of the BTC (first moment). This allows for the calculation of a partition coefficient Kd for instantaneous sorption. Estimated values of Kd were 1.4 and 1.3 L kg-1 from CDeq, and 1.5 and 1.7 L kg-1 from CDnoneq.
In the two-site model, the nonequilibrium dimensionless parameters ß and were used for calculation of the fraction of Type 1 sites, f, and the first-order rate (d-1) for the kinetic, Type 2 sites. The estimated values for the parameters f and were different for the fits with or without degradation term. If degradation is neglected, then the fraction of sites at equilibrium
would indicate that the majority of the sites are of the kinetic type, whereas if a sink term is introduced, then the proportion of kinetic sites would be much lower (57%). The estimates of the first-order rates for the sorption kinetics are extremely high (>58 d-1), and they increase if degradation is not accounted for.
Mass balance of the amount of triticonazole applied was obtained from the total amount of radioactivity recovered from the columns, and this showed that 95% of the total amount was detected in the effluent. Nevertheless, results of the extraction and analysis of the soil sections at the end of the experiments showed that a fraction of 5.3% was recovered as methanol-extractable (2.7% of total applied) and bound 14C residues (2.6% of total applied). The distribution of the extractable and bound 14C residues in the soil profiles (Fig. 6)
show that the amount of extractable residues increased with depth in the column, whereas the bound residues were mainly located in the upper end of the column (at the inlet), and they decreased with depth.
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Fig. 6 Measured concentration profiles of the extractable (solid line) and bound (dotted line and symbols) 14C residues remaining in the soil columns after elution of a pulse of 0.5 pore volume of 14C-triticonazole at 5 mg L-1 with 16 pore volumes of water. Total recovered amounts of the extractable and bound fractions respectively averaged 2.7 and 2.5% of the initial amount applied
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Discussion
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Flow Characteristics in Grignon Soil
Symmetrical, nonretarded BTC of chloride are expected for the displacement of a nonreactive tracer in repacked homogeneous water-saturated columns of sieved soils, as reported, for instance, by Gamerdinger et al. (1990) and Chen and Wagenet (1997). The invariant, very low dispersivity measured for Grignon soil is smaller than the aggregate size of the sieved soil, which suggests that some destruction of the aggregates occurred during packing of the saturated soil columns. Low dispersivities and high Peclet numbers (P) indicating poor dispersion are characteristically reported in well packed homogeneous soil columns (Lee et al., 1988; Romero et al., 1997). In undisturbed soils, where preferential flow may occur, much higher dispersivities, ranging from 4.5 to 65.8 cm, were observed by Jaynes (1991), and a low P of 0.7 was reported by O'Dell et al. (1992).
The adequate fit of CDeq to our experimental chloride data and the symmetrical shape of the BTC at the three measured water velocities in our study is representative of systems that are not influenced by transport nonequilibrium. Thus, physical nonequilibrium, which is suggested by the small decrease in ß that was observed, can reasonably be assumed to be negligible in our columns, and nonequilibrium conditions for triticonazole reactive tracer would mostly arise from sorption nonequilibrium.
Nonequilibrium Transport of Triticonazole
The asymmetrical shape of the triticonazole BTC is indicative of nonequilibrium, and the considerable tailing suggests that it is primarily sorption related. Similar sorption-related nonequilibrium transport characteristics have been reported for various organic chemicals in repacked homogeneous soil columns (Lee et al., 1988; Angley et al., 1992; Gaber et al. 1995), and in field studies (Jaynes, 1991). The failure of the CDeq model is then expected, since the LEA would not be valid for triticonazole transport in soil.
Sorption nonequilibrium with a two-stage approach to equilibrium has been evidenced for a large number of pesticides. A short, initial fast phase of sorption is generally reported in the first few minutes (Kookana et al., 1993; Gaston and Locke, 1995), which is followed by an extended period of much slower uptake, occuring over periods of days or months. Brusseau and Rao (1989), and Brusseau (1991) reviewed the different rate-limiting processes of nonequilibrium, and they attributed the rate-limiting step in the sorption nonequilibrium of HOC in soil to intrasorbent (intraorganic matter) mass transfer diffusion. Our previous results (Beigel et al., 1997 and 1999) support this hypothesis, because triticonazole sorption was related to nonspecific interaction mechanisms with soil OM, and because rate-limiting, slow desorption of triticonazole was evident during prolonged incubation, which was attributed to intrasorbent diffusion. Under these conditions, instantaneous Type 1 sites would represent the sites that are directly accessible, while kinetic Type 2 sites would be sites that are more remote in the soil organic constituents. The first-order rate for sorption kinetics of the time-dependent sites would in fact reflect a rate-limiting diffusion process. If the water velocity is low enough for all the accessible Type 1 sites to be reached, the two-site assumption may be valid, and this approach has been successfully used to predict the nonequilibrium transport of pesticides in homogeneous soil columns under low velocities (Gamerdinger et al., 1990; Gaber et al. 1995).
The CDnoneq model correctly described the early tailing of the asymmetrical distal portion of triticonazole elution peak, but it failed to describe the extended tailing of triticonazole BTC. The two-site approach is thus not adequate for predicting the transport of triticonazole in soil at high velocity. Chen and Wagenet (1997) observed similar failure of a two-site, first-order model to describe atrazine [6-chloro-N-ethyl-N'-(1-methyl-ethyl)-1,3,5-triazine-2,4-diamine] transport in homogeneous soil columns at high velocities. In such cases, the rate of sorption–desorption from the kinetic sites cannot be described by a single first-order rate constant, as also reported by Connaughton et al. (1993). The two-stage tailing that we observed suggests that at least two types of kinetic sites need to be considered in addition to the instantaneous sorption sites.
The estimated Kd values in the range of 1.3 to 1.7 L kg-1 account for the instantaneous sorption of triticonazole during transport. These estimates were much lower than the measured batch equilibrium Kd of triticonazole of 4.35 L kg-1, indicating that the batch equilibrium Kd would considerably overpredict the retardation of triticonazole during transport at high flow rate. Such leftward shift of the experimental BTC compared with batch-measured partition coefficient predictions has been reported frequently at high velocities, while better agreement is obtained at lower velocities (O'Dell et al., 1992; Gaber et al., 1995; Chen and Wagenet, 1997). This clearly shows that the batch-measured partition coefficient would not be appropriate for description of the transport of triticonazole when the residence time in soil is not long enough for some sorption sites to achieve equilibrium.
The very high value of the first-order rate (>58 d-1) estimated from the two-site model, which accounts for the asymmetry in the elution peak, indicates a quick transport to accessible sorption sites. In the water-saturated soil columns packed with finely sieved soil, triticonazole is directly in contact with the dispersed soil organic constituents, and it can then rapidly diffuse to accessible sorption sites in the internal voids of organic matter. A great portion of the kinetic sorption sites would, thus, rapidly attain equilibrium and extensively increase the sorption of triticonazole.
On the other hand, the extended slow and continuous release of sorbed residues and the fraction of extractable residues remaining in the column suggest that another part of the kinetic sites would be much more rate-limited. Methanol-extractable residues may be subject to desorption in water, as observed by Barriuso et al. (1992), and thus are potentially available for transport. Indeed, their location at the effluent end of the soil indicates that they were subject to convective–dispersive transport, and their release would continue after the 16 pore volumes of water were applied, resulting in a longer tailing than observed. These slow kinetics may be attributed to rate-limiting sorption–desorption that is controlled by mass transfer diffusion from remote soil sites. This was not accounted for in the two-site model, and it needs to be determined independently for an adequate description of triticonazole transport in soil. Use of the batch equilibrium Kd would not be appropriate, as the batch-obtained value does not represent the true sorption equilibrium of triticonazole in soil (Beigel et al., 1997). Measurement of apparent desorption coefficients (Kapp) at prolonged incubation times may be more adequate for evaluation of the time-dependent desorption, as shown by O'Dell et al. (1992) and Beigel et al. (1997).
Kinetics of Bound Residue Formation
In our previous experiments (Beigel et al., 1997 and 1998), the formation of a significant fraction of soil-bound residues of triticonazole in both nonsterile and 
-radiated sterile soil was evident immediately after treatment. In such conditions, bound residue formation may arise from physical trapping in the internal voids of soil organic matter (Khan, 1982; Calderbank, 1989). Bound residues are resistant to desorption and would not be available for transport. This would affect the elution of triticonazole in soil and might also partly account for the two-site, two-region model's incapacity to describe the experimental data. Indeed, small but significant amounts of bound residues were detected in the soil columns. Their localization at the inlet end of the soil profile shows that the removal from soil solution and stabilization in the soil colloids occurred immediately after treatment and that those residues were not subject to convective transport. The nonreversible disappearence of triticonazole from soil solution as bound residues cannot be accounted for by sorption in the models, even in the two-site approach, which explains why the fraction of kinetic sites is considerably higher when no sink term is provided. Bound residues would be partly accounted for in the model by the degradation term, and the relatively high degradation rate µliq that is estimated from the two-site model may be attributed to the rapid formation of triticonazole-bound residues. Hence, the estimated decay rate would depend on the residence time in the columns (i.e., on the experimental set up) and is, therefore, unreliable. Moreover, the rapid formation of bound residues in soil is followed by a much slower, continuous stabilization (Beigel et al., 1999), and a single first-order rate would not be suitable to describe the dissipation of triticonazole as stabilized, bound residues. The kinetics of bound residue formation need to be further studied and correctly described to allow for an accurate modeling and prediction of transport in soil.
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Conclusion
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Our results clearly show that triticonazole convective–dispersive transport in homogeneous saturated soil columns is primarily influenced by its interactions with the soil matrix through rate-limited sorption–desorption and through the formation of bound residues. Both processes may be related to rate-limiting, intrasorbent diffusion into soil organic constituents and thus would be strongly dependent upon the water flow rate. Failure of the two-site, nonequilibrium CDE model to describe the extended tailing of triticonazole experimental BTC at a high flow rate shows that nonequilibrium sorption cannot be accounted for by a single first-order rate constant. Similarily, whereas the immediate formation of bound residues may be accounted for with a high first-order rate in the decay term, use of a single rate constant would prove erroneous for describing the much slower, continuous stabilization of triticonazole residues during prolonged incubation time, which has been evident in previous studies.
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ACKNOWLEDGMENTS
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This research has been conducted under the Bio Avenir programme funded by Rhône-Poulenc with the participation of the French Ministère de la Recherche et de l'Espace and Ministère de l'Industrie et du Commerce Extérieur. All the radiochemical analyses were performed at Laboratory of Soil Science, INRA, 78850 Thiverval-Grignon, France. The authors wish to thank Ghislain Sevenier for performing the Cl- analysis.
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NOTES
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1 Trade names and company names are included for the benefit of the reader and do not imply endorsement or preferential treatment of the product listed by INRA. 
Received for publication June 30, 1998.
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INTRODUCTION
