A Distributed Reactivity Model for Sorption by Soils and Sediments. 4

Environ. Sci. Technol. 1996, 30, 881-888

A Distributed Reactivity Model for Sorption by Soils and Sediments. 4. Intraparticle Heterogeneity and Phase-Distribution Relationships under Nonequilibrium Conditions WALTER J. WEBER, JR.* AND WEILIN HUANG Environmental and Water Resources Engineering, Department of Civil and Environmental Engineering, The University of Michigan, Ann Arbor, Michigan 48109-2125

Rates of phenanthrene sorption by four different types of soils and sediments were characterized by examining the time dependence of solute phase distribution relationships (PDRs) in completely-mixed batch reactors. Unlike conventional single-level concentration methods, the experiments were conducted using a range of concentrations to obtain a time series of nonequilibrium PDRs for each sorbent-sorbate system over reaction periods ranging from 1 min to 14 days. In all cases tested, the nonequilibrium PDRs changed from approximately linear form to increasingly nonlinear form as the time of reaction increased. A Freundlich-type relationship, q(t) ) KF(t)C(t)n(t), was used to relate values of measured temporal solid-phase solute concentrations, q(t), to corresponding solution-phase solute concentrations C(t). After a short “initiation” stage, the parameters n(t) and KF(t) were observed to vary functionally with logarithmic time. A three-domain particle-scale model predicated on the existence of discrete soil components (exposed inorganic surfaces and amorphous and condensed soil organic matter) is invoked to explain the observed sorption behavior and functional relationships underlying the time dependence of the PDRs.

Introduction Because of its effects on sorption and desorption patterns associated with hydrophobic organic contaminants (HOCs), soil and sediment heterogeneity exerts a significant influence on the efficiency and effectiveness of various site remediation methods. Heterogeneity at both microscopic and macroscopic scales has been shown to affect dominant * Corresponding author e-mail address: [email protected]; telephone: 313-763-1464; fax: 313-763-2275.

0013-936X/96/0930-0881$12.00/0

 1996 American Chemical Society

sorption mechanisms, isotherm nonlinearity, solute-solute competition, and binding energies associated with the interactions of HOCs with soils, sediments, and other environmental sorbents (1-6). This paper examines sorbent heterogeneity at the particle scale, specifically with respect to its effects on sorption rates and equilibria. Traditional experimental approaches to the investigation of sorption rates center on time-series observations of residual solute concentration in a single system initiated at a single level of concentration (7-22). We employ an approach in which solute phase distribution relationships (PDRs) obtained in experimental systems initiated at different starting concentrations are measured as functions of time. A PDR comprises a multiple-concentration measurement of the distribution of a solute between an aqueous phase and a solid phase at a given time during the course of a sorption or desorption process; the ultimate PDR for a system is that obtained for a condition of equilibrium (i.e., a sorption isotherm). We then relate observed time-dependent changes in the characteristics of the PDRs to the sorption properties of soil particle components to delineate the influence of particle-scale heterogeneity on overall sorption rates. We have previously introduced the concept of distributed reactivity to describe adsorption equilibria in soil and sediment systems (2). In the context of the distributed reactivity model (DRM), sorbate molecules access different regions or domains of a particle via different sorption mechanisms. If the concept is valid, it is further reasonable to expect that sorbate molecules also access different particle domains at different times and rates as sorbent loading proceeds. The dynamics of sorption processes of this nature cannot be characterized readily using conventional methods. We have therefore devised an alternative experimental method for observing rates of HOC sorption by sorbents that exhibit intraparticle heterogeneity. We then relate discrete demarcations in the rate of PDR parameter change to changes in the sequence of solute access to different sorption domains. Previous papers in the DRM series have advanced the concept and demonstrated the fact that soils and sediments can be treated as heterogeneous combinations of active organic and inorganic components with respect to sorption equilibria, each component having its own sorption energy and sorptive properties (2-4). The DRM is predicated on the assumption that the overall sorption isotherm for a natural solid is the sum of the sorption isotherms of the active components, each component exhibiting either a nonlinear (e.g., Freundlich or Langmuir) or a linear sorption with respect to a particular solute. This paper extends the DRM to sorption rate processes for heterogeneous natural solids under nonequilibrium conditions. Multiple experimental systems containing various initial aqueous-phase solute concentrations are used to measure rates of uptake under different gradients in solute concentration. Sorption data measured at a given time thus include a set of corresponding aqueous-phase and solid-phase concentrations, each set comprising a sorptive phase distribution relationship at a specific time.

VOL. 30, NO. 3, 1996 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 1

Soil and Sediment Characteristics sample EPA-23 (sediment) EPA-22 (sediment) EPA-20 (soil) EPA-15 (sediment) a

TOC (wt %)

N2-BET surface sanda silta claya (wt %) (wt %) (wt %) area (m2/g)

2.35 ( 0.29 1.69 ( 0.21

33.2 3.25

17.3

13.6

69.1

26.1

52.7

21.2

1.26 ( 0.17

10.2

0.0

71.4

28.6

0.97 ( 0.19

15.2

15.6

48.7

35.7

Based on data from ref 24.

Experimental Section Sorption experiments were conducted at room temperature (23 ( 2 °C) using completely-mixed batch reactors (CMBRs). Fixed sorbent dosages and a range of initial aqueous-phase solute concentrations were employed to obtain time-series measurements with which to calculate corresponding PDRs. Reaction times varied from 1 min to 14 days. Each point on each PDR was determined from an independent experiment, each conducted at a different initial solute concentration in an individual CMBR. Each PDR involved at least 12 such experiments, spanning approximately 2 orders of magnitude in aqueous-phase solute concentration. Sorbents and Their Characterization. Four EPA soil and sediment samples (EPA-15, EPA-20, EPA-22, and EPA23) obtained from W. L. Banwart of the Department of Agronomy at the University of Illinois at ChampaignUrbana were used as natural sorbents. A number of different soil and sediment samples have been employed in the work we described in the three earlier papers in this series (2-4). For this work, sorbent selections were extended to other soils and sediments to more broadly examine the effects of soil heterogeneities on sorption processes. The solids selected were collected in 1980 and have been well characterized (23, 24). Importantly, they have also been employed in other studies of organic contaminant sorption under both equilibrium and nonequilibrium conditions (7, 23-29). A standard soil splitting procedure was used to split the 500-g samples received into smaller subsamples to ensure that they were representative of the original sample. The subsamples were stored in air-tight containers until used in the experiments. The soil and sediment samples were characterized, and the results were compared to those reported earlier (23, 24) to determine whether their major properties had changed over the 15 years since they were collected. TOC was analyzed using a high-temperature combustion method (CHN-1000 analyzer, Leco, Inc.) on samples pretreated with a dilute phosphoric acid solution (30). Specific surface areas and microporosities were evaluated using techniques suggested by Gregg and Sing (31) and 90-point N2 adsorption-desorption isotherm data collected for each sample at liquid nitrogen temperature (ASAP-1000, Micromeritics). Major sorbent characteristics are listed in Table 1. TOC analyses show little change in the organic carbon contents of the samples after their collection. Particle size distribution analyses and specific surface areas revealed that the majority of the particles are of clay and silt size. Pore analyses suggested that no significant intraparticle porosity existed in these soils and sediments.

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ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 30, NO. 3, 1996

Chemicals and Solution Preparation. Phenanthrene, the solute employed in the studies reported here, was obtained in spectrophotometric grade (98%) from Aldrich Chemical Co., Inc. A primary stock solution was prepared by dissolving an appropriate amount of solute in methanol (HPLC grade) and sequentially diluting with methanol to make a series of stock solutions of various concentrations. All stock solutions were stored at 4 °C in glass bottles sealed with Teflon-lined tops. A desired amount of stock solution was mixed with a background solution in a volumetric flask to make an initial aqueous solution for the sorption experiments. CaCl2 at a level of 0.005 M was the major mineral constituent of the background solution, and 100 mg/L NaN3 was added to control biological activity. The solution was buffered at pH 7 with 5 mg/L NaHCO3. Phenanthrene concentrations in these initial aqueous solutions were analyzed by reverse-phase HPLC, as described in detail below. Stock concentrations and volumes transferred were designed to maintain a methanol concentration of 0.2% by volume in the initial aqueous solutions, a level at which methanol has no measurable effect on sorption (32). Sorption Experiments. The CMBRs consisted of glass centrifuge bottles (25 mL, Corex) sealed with screw caps, Teflon-lined silicone septa, and silver foil; the latter was used to line the Teflon septa to minimize the loss of phenanthrene to reactor compartments. Each reactor contained approximately 0.1 g of sorbent and 30 mL of aqueous solution. Reactors used for PDR measurements at times >1 h were filled to a minimum headspace with aqueous phenanthrene solution and immediately weighed and placed in a rotary tumbler for complete mixing at 12 rpm for the predesigned contact time. Reactors used for PDR measurements at times e1 h were filled with a predetermined amount of aqueous phenanthrene solution and mixed at the centrifugation site using a bench-top tumbler operated at the same rotary speed. The latter procedure was used to minimize errors in reaction time associated with short-term PDR measurements. Solids were separated from the aqueous solutions by centrifugation as soon as the reactors were taken from the tumblers. Reactors used for PDR measurements at times e1 h were centrifuged at 10 000 rpm (7800G) for 4 min, again to reduce errors in reaction time, while reactors used for PDR measurements at times >1 h were centrifuged at 8500 rpm (5600G) for 10 min; no differences were found between the two centrifugation techniques. Supernatant was immediately withdrawn from each reactor after centrifugation. A preliminary study had shown that, depending on solute concentration, about 1-7% of the aqueous-phase phenanthrene was lost by sorption to the glass walls of the GC vials during analysis. To eliminate this loss, the supernatant was mixed with about 2 mL of methanol in a 5-mL vial capped with a Teflon top. The amounts of supernatant and methanol were weighed, and the dilution factor for the supernatant was calculated from density data for mixtures of water and methanol (33). The dilution factor was then multiplied by the phenanthrene concentration analyzed for the mixture to obtain the concentration in the equilibrated aqueous solution. The result obtained by this procedure is equivalent to analysis of a 1:1 (volume) liquid-liquid hexane extract of the same aqueous solution. The phenanthrene concentration in the mixture was analyzed using reverse-phase HPLC (ODS, 5 mm, 2.1 × 250 mm column on a Hewlett-Packard Model

TABLE 2

Equilibrium Parameters for Sorption of Phenanthrene Freundlich model

av KDa values for different Ce (µg/L) ranges

linear model

sample

log KFa

n

Nc

R2

KD

R2

EPA-23 EPA-22 EPA-20 EPA-15

0.235 -0.137 -0.191 -0.058

0.727 0.890 0.756 0.726

18 43 18 20

0.995 0.996 0.994 0.993

0.38 ( 0.02 0.29 ( 0.01 0.17 ( 0.01 0.15 ( 0.01

0.948 0.992 0.979 0.970

100

KD

differenced

0.67 (6) 0.34 (13) 0.27 (5) 0.30 (6)

0.37 (3) 0.28 (16) 0.17 (6) 0.17 (6)

0.286 0.201 0.156 0.114

23 28 8 33

a K in (µg/g)/(µg/L)n; K in (µg/g)/(µg/L). b Calculated from K and f data reported by Karickhoff (26) for phenanthrene and the same set of soils F D oc oc and sediments. c Number of observations. d % difference ) [(measured KD (Ce > 100 µg/L) - reported KD)/measured KD (Ce > 100 µg/L)] × 100.

1090) with a diode array UV detector (250-nm wavelength) for concentrations ranging from 20 to 1000 µg/L and a fluorescence detector (Model HP 1046A, 250-nm excitation wavelength, 364-nm emission wavelength) for concentrations ranging from 0.5 to 50 µg/L. External phenanthrene standards (in methanol) were used to establish both a linear calibration curve for the UV detector and a second-order exponential calibration curve for the fluorescence detector. The elute solvent used was a mixture of acetonitrile (90%) and water (10%). The solid-phase concentration of phenanthrene was computed based on a mass balance of solute between the two phases. Control experiments in reactors containing no sorbent were also run to assess loss of solute to reactor components. Average system losses were shown to be consistently less than 3% of initial solute concentration, and no correction was required. Reaction Time Error. The relative error in reaction time reported for each PDR was estimated based on the absolute error, which was about the same for all experiments. Reaction time is defined as the time measured from completion of reactor filling to the end of tumbling, while the actual sorbent-solution contact time is the time from the beginning of reactor filling to the beginning of centrifugation at full speed. The majority of the suspended particles should have been deposited on the bottom of the reactor once the rotor achieved full speed. The absolute error is a function of the difference between the reaction time and the sorbent-solution contact time. The approximate relative error was about +67% for 1 min, +20% for 5 min, +10% for 10 min, +4% for 30 min, and +2% for 1 h of mixing time. The error is negligible when the reaction time is greater than 1 h.

Results The linear and Freundlich sorption isotherm models are given respectively by

qe ) KDCe

(1)

qe ) KFCen

(2)

and

where qe and Ce are the equilibrium solid-phase and aqueous-phase solute concentrations, respectively. Both models were tested with respect to goodness of fit to the equilibrium sorption data for all four EPA soil and sediment samples; data used to fit the Freundlich model were logtransformed. The results, along with standard deviations, measured average KD values in three different aqueous equilibrium solute concentration ranges, and calculated KD values based on a Koc-Kow correlation reported by

Karickhoff (26) for the same samples are listed in Table 2. The sorption data used in Karickhoff’s correlation were measured using batch systems mixed for 2 days at concentrations approaching the aqueous solubilities of phenanthrene and the other solutes studied (26, 34). It is clear from Table 2 that all of the isotherms are significantly nonlinear, as evidenced by the Freundlich n values, and that the Freundlich model fits the equilibrium data better than the linear model for all samples examined. The average KD values obtained in this study decrease appreciably as equilibrium aqueous-phase solute concentrations increase and are all consistently higher than the corresponding values calculated from the Karickhoff correlation, confirming nonlinear behavior over the wider solute concentration ranges employed in our studies. These findings challenge the assumption of linear partitioning made by other investigators (7, 23-29). Further comparison of KD values between the two studies shows that the percentage difference in the last column of Table 2 is approximately proportional to the weight percentages of the sand fraction in these samples (Table 1), implying that the sorption data used by Karickhoff to establish the KocKow correlation in question was not equilibrium data. Solute phase distribution relationships (PDRs) reported here were obtained by fitting eq 2 to log-normalized nonequilibrium sorption data, using a quasi-Newton method (SYSTAT) to obtain the PDR parameters n(t) and KF(t) within confidence levels of 95%. The regression results yielded R2 values equal to or greater than 0.99, confirming visual observations that the Freundlich-type model provided an excellent fit of the nonequilibrium data for all systems studied. A typical series of PDRs for sorption of phenanthrene by EPA-23 sediment is illustrated in Figure 1. A significant feature observed for this PDR time series, and for the time series for all of the other sorbents as well, is that the PDR is only modestly nonlinear at the shortest reaction time measured but becomes increasingly nonlinear as the reaction proceeds. The statistical best fit of eq 2 to the 1-min PDR data in Figure 1 has an n(t) value of n(1 min) ) 0.93 ( 0.03. For a reaction time of 1 h, n(t) decreases to n(1 h) ) 0.80 ( 0.03, and at 1 day reaches a constant value of n(1 d) ) 0.73 ( 0.03, which corresponds to the apparent equilibrium value of n for this system. Along with decreasing linearity, the sorption capacity, KF(t), increases sharply from 0.11 ( 0.01 at 1 min to 0.58 ( 0.06 at 1 h and to 1.37 ( 0.08 at 1 day, the latter corresponding to about 81.5% of the value attained at the reaction time required to reach an apparent equilibrium (14 days). This pattern of rapid early uptake of solute followed by a slow approach to equilibrium is similar to that observed by investigators using single starting concentrations (7).


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FIGURE 4. Changes in PDR coefficients for sorption of phenanthrene by EPA-15 sediment as a function of log time. Error bars represent (2σ (95% confidence levels).

FIGURE 1. Time-dependent PDRs for sorption of phenanthrene by EPA-23 sediment.

FIGURE 5. Changes in PDR coefficients for sorption of phenanthrene by EPA-20 soil as a function of log time. Error bars represent (2σ (95% confidence levels). FIGURE 2. Changes in PDR coefficients for sorption of phenanthrene by EPA-23 sediment as a function of time. Error bars represent (2σ (95% confidence levels).

FIGURE 6. Changes in PDR coefficients for sorption of phenanthrene by EPA-22 sediment as a function of log time. Error bars represent (2σ (95% confidence levels). FIGURE 3. Changes in PDR coefficients for sorption of phenanthrene by EPA-23 sediment as a function of log time. Error bars represent (2σ (95% confidence levels).

To avoid crowding of the figure, only four of the 11 PDRs measured for EPA-23 sediment are shown in Figure 1. The PDR parameters, KF(t) and n(t), for all 11 studies are shown in Figure 2 as a function of time, t, and in Figure 3 as a function of log t. Similar log t plots are given in Figures 4-6 for EPA-15, -20, and -22 soils and sediments, respec-

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tively. Close inspection of these figures reveals that changes in the parameters appear to occur in three stages: (i) an initiation stage; (ii) a logarithmic stage; and (iii) an apparent equilibrium stage. The term initiation stage references a period over which the PDR has a linear or near-linear form. EPA-20 soil (Figure 5) and EPA-22 sediment (Figure 6) appear to have initiation stages of approximately 10 min, while EPA-23 (Figure 3) and EPA-15 (Figure 4) sediments exhibit little if any initiation period. In the logarithmic stage, the PDR n(t) values decrease as a function of

FIGURE 7. Schematic illustration of domain types associated with a soil or sediment particle.

logarithmic time, while the KF(t) values increase similarly. A simple regression of the PDR parameters for EPA-23 soil over logarithmic times gives the following correlations between n(t) and time (t) and KF(t) and time (t) within their respective logarithmic stages:

n(t) ) 0.93 - 7.25 × 10-2 log t for t ) 0-240 min (R2 ) 0.990) (3) KF(t) ) -3.49 × 10-1 + 5.377 × 10-1 log t for t ) 30-2880 min (R2 ) 0.993) (4) Standard deviations for the first coefficient in eqs 3 and 4 are (0.10 × 10-1 and (0.53 × 10-1, respectively, while those for the respective prelogarithmic coefficients of these equations are (0.32 × 10-2 and (0.21 × 10-1. These standard deviations might be larger if the errors in PDR parameters and reaction times were considered during the regression. The sharpened rate of change of the parameters n(t) and KF(t) that occurs once the initiation stage ends and the logarithmic stage begins suggests that the solute has begun to access a new domain on or within the solid phase. This domain appears to be energetically more heterogeneous and to have higher sorption capacities for phenanthrene, as evidenced respectively by a rapidly decreasing n(t) and a rapidly increasing KF(t) . After about 1 day, both n(t) and KF(t) begin to transition to their respective apparent equilibrium stage values. The parameter n(t) achieves an essentially constant value once sorbate molecules have accessed all energetically different sorbent domains. The sorption process then continues beyond this point until no further changes in the parameter KF(t) are measurable, a condition which may or may not represent a true thermodynamic equilibrium.

Discussion The time dependence of the form and characteristics of the PDRs observed in this work provides insight to changes in dominant sorption mechanisms over time, which in turn can be related to intraparticle heterogeneity. Particular associations of the active components of a soil or sediment particle determine the sequence of accessibility of each of those components, and it is this sequence that we feel is reflected by changes in PDRs over time. Figure 7 characterizes schematically a three-domain model involving multiple sorption reactions at the particle scale that we

propose by way of interpretation of the time-dependent PDR patterns observed in this study. This three-domain model comprises (i) an exposed mineral domain, (ii) an amorphous soil organic matter domain, and (iii) a condensed soil organic matter domain. The contributions of unexposed mineral surfaces to overall sorption are considered negligible in our model. Each domain exhibits somewhat different characteristics and behavior with respect to the energetics, rates, and mechanisms associated with HOC sorption. The model is based on the hypothesis that soil organic matter (SOM) can be treated as two different classes of macromolecular aggregates, an amorphous class and a condensed class, a hypothesis suggested in several recent investigations (2-4, 35). We further hypothesize that the nonlinearity of a PDR relates principally to the microscopic heterogeneity of the condensed SOM and that changes in the characteristics of a PDR with time for a given system are associated with a particular sequence of domains. There is strong evidence in the literature of geochemistry and soil sciences to support the co-existence of the three domains pictured in Figure 7. The existence of relatively impervious mineral surfaces of domain I type in soil and sediment particles is well known. Support for the existence of domain II is found in two types of recent studies. The first of these has shown that fulvic and humic acids are sorbed in essentially monolayer form on the mineral surfaces of soils and sediments (36-38). The second has revealed that humic and fulvic materials dissolved in aqueous solution have a tendency to form micelle-like structures, with hydrophilic shells facing the aqueous phase and hydrophobic cores in the center of the aggregates (39). This self-organized structure acts to enhance the solubility of HOCs in such solutions by allowing them to penetrate the hydrophilic shells and become trapped in the hydrophobic core of the macromolecular organic aggregates. It has been suggested that part of the mineral-bound SOM may retain the amorphous form it exhibits in solution and have a tendency to swell when interacting with water (4, 35). Wershaw has proposed a membrane-like bilayer model to describe the structure and organization of humic SOMs bound on mineral surfaces and has used this model to explain their related partitioning-like behavior with respect to sorption of HOCs (40-42). It is further widely recognized in organic geochemistry that diagenetic processes act over geological times to gradually increase the degree of condensation and crystallinity of soil- and sediment-associated SOMs, along with their C/O and C/H atomic ratios (43-45). We postulate that domain III of the model given in Figure 7 contains such condensed SOMs, perhaps even exhibiting a certain degree of microcrystallinity. Fulvic and humic acids extracted from soils have shown the existence of crystallized regions having X-ray diffraction spectra with two broad peaks, one near 0.35 nm and the other in the 0.41-0.47-nm region (46, 47). These peaks are often cited as evidence of tightly packed aromatic sheets (0.35 nm) and less tightly packed aliphatic chains (0.41-0.47-nm). Sorption of HOCs by condensed SOM appears to be energetically more favorable and more nonlinear than that by amorphous SOM, most likely because of the more heterogeneous composition and less polar character of condensed organic matter (4, 5, 43-45). Moreover, sorption by condensed SOM appears to be very slow, and this we attribute to slow intradomain diffusion (4, 35, 49, 52, 53).


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FIGURE 8. Distributed reactivities and sequencing of sorption processes pursuant to the three-domain model.

The time span of the logarithmic stage of n(t) (e.g., Figure 3) suggests that it takes approximately 1-2 days for sorbate molecules to access representative condensed SOM domains associated with the soil and sediment aggregates studied. These domains are probably located in the outer regions of soil aggregates or associated with the finest particles of the bulk soil sample. The fact that KF(t) has a much longer logarithmic stage than n(t) indicates that sorption by the remaining condensed SOM in domain III requires a much longer equilibration period. The time required to achieve overall sorption equilibrium is thus a function of soil particle size and the properties of pores connecting intraparticle condensed SOMs with the bulk solution. Our observation that n(t) decreases sharply as a function of time suggests that sorption from solutions of higher initial solute concentration approaches equilibrium faster than does that initiated at lower solute concentrations. This would necessarily impose a concentration dependence on the coefficient of any first-order (reaction or diffusion) rate model employed to interpret the data, suggesting that the process associated with domain III is in fact nonlinear or non-first-order; i.e., the rate is a function not only of aqueous solute concentration but also of the properties of the condensed SOM as well. In other words, the observed rate behavior is fundamentally and phenomenologically inconsistent with any existing mathematical model that is based on assumptions of first-order (reaction and/or diffusion) processes (see, for example, refs 11-17, 35, 48, and 49). The nonlinear sorption rate behavior we observe might be best modeled either by invoking a Langmuir-type kinetic model, which is first order with respect to both aqueous solute concentration and the density of empty sorbent sites, or by employing a non-Fickian diffusion model in which the coefficient of sorbate diffusion is concentration dependent (50, 51). While sorptions in condensed SOM domains are relatively slow and nonlinear, sorptions associated with mineral surfaces and amorphous SOM are effected by simple linear and relatively fast processes. We suggest that these

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processes dominate the early near-linear PDRs observed in this work. Although HOC uptake by mineral surfaces has been shown to be insignificant when the foc of a soil is greater than 0.1 wt % (54, 55), it is an important conceptual domain in our model. In general, uptake of HOCs by mineral surfaces is fundamentally an adsorption process, one that likely generates a localized Langmuir-type PDR. Because both sorption energies at mineral sites and solution-phase concentrations are generally low, surface coverage is very small in most cases, and probably limited to Henry’s range of the generalized PDR. For practical purposes, then, we can expect domain I sorption to be near-linear and to exhibit pseudo-first-order rates, the latter because the rate-controlling mechanism associated with this domain is likely either adsorption without site limitations or boundary-layer mass transfer, both of which are linear functions of solution-phase concentration (first order) as well as relatively rapid processes (56). We hypothesize that sorption in domain II is largely a partitioning process and thus can be described by a linear local PDR model. Sorption sites in this highly hydrated, loosely knit, flexible, and amorphous SOM domain are readily accessed by HOC molecules, and associated sorption rates are thus relatively fast. To give a frame of reference, sorption of PAHs by dissolved fulvic and humic acids is typically complete in about 10 min (57-60). Once the readily accessible domain I and domain II sites have reached apparent equilibrium, the overall sorption process is dominated by domain III behavior. This point corresponds to the onset of the logarithmic stage, during which the characteristics of the PDR, as reflected by both n(t) and KF(t), begin to change dramatically. Once domain III has been accessed, the coefficient n becomes essentially constant, suggesting that the heterogeneous characteristics of the condensed SOMs in these particular samples are reasonably uniform within the domain and are probably independent of soil particle size and independent of the location of the domain within an individual particle. Sorption processes in domain III and at internal domain I and domain II sites trapped in larger particles continue after this point, and KF(t) therefore continues to increase until complete equilibrium between the sorbent particles of various sizes and the solution phase is obtained. The time required for sorption to approach this equilibrium is thus highly dependent on soil particle size and association patterns of the three domains at a particle scale. A schematic representation summarizing the “process dynamics” associated with our three-domain model is given in Figure 8, along with characterization of the corresponding time-dependent PDR coefficients n(t) and KF(t). This schematic represents the sequence in which sorbate molecules might access a range of heterogeneous sorption sites distributed in a soil or sediment particle. Sorption process occurring in each domain may comprise a range of parallel sorptive reactions having different rate coefficients and even different mechanisms. As discussed above and pictured in Figure 8, we suggest that sorption processes begin simultaneously in domains I and II and that the corresponding global PDR is thus the linear sum of two linear local PDRs. The initiation stage corresponds to the time required for migration of solute molecules through domain II to access domain III. The global PDRs corresponding to this stage thus change only slowly from their initial linear form. The relatively small deviations from linearity observed during this stage most likely relate to

differences in the times required for solute molecules to penetrate amorphous SOM layers of different thickness in order to begin interacting with condensed SOMs. Some linear sorption can continue beyond the initiation stage due to slow intraparticle transport to mineral and SOM sites internal to the particle that have contact with the aqueous phase, but the amount of prolonged sorption attributable to this uptake is likely small. The experimental observations upon which we base our three-domain model are, as noted above, phenomenologically different from what any model based on assumptions of first-order processes (reaction and/or diffusion) would predict. A model that incorporates a nonlinear equilibrium relationship and one or more first-order reaction rate or diffusion relationships will generate a simple X-type pattern of change in n(t) and KF(t), rather than the more complex pattern illustrated in Figure 8. Such X-type patterns exhibit no initiation stages for n(t) and KF(t), both of which parameters also arrive at their respective equilibrium values at the same time. Figure 8 shows initiation stages for both parameters as well as a region in which KF(t) continues to increase dramatically after n(t) has become essentially constant. We argue that these phenomenologic differences reflect fundamental differences between the three domain model and models based on simple first-order reaction rates and/or diffusion processes.

Acknowledgments The authors thank Hong Yu, Joel Rogers, and Christina Gibbs for their assistance in the experimental phases of this work and Dr. Wayne Banwart of the Department of Agronomy at the University of Illinois at Urbana-Champaign for providing the EPA soil and sediment samples. We appreciate also the helpful feedback and insights contributed by Dr. Mark A. Schlautman, a postdoctoral research fellow in our program at the time of this work and currently a faculty member in Civil and Environmental Engineering at Texas A&M University. The research presented here was funded by the U.S. Environmental Protection Agency, Office of Research and Development, in part by Cooperative Agreement CR818213-01-0 with the Risk Reduction Engineering Laboratory, Cincinnati, OH, and in part by Grant R-819605 to the Great Lakes and Mid-Atlantic Center (GLMAC) for Hazardous Substance Research. Partial funding of the research activities of GLMAC is also provided by the State of Michigan Department of Natural Resources.

Literature Cited (1) Weber, W. J., Jr.; McGinley, P. M.; Katz, L. E. Water Res. 1991, 25, 499-528. (2) Weber, W. J., Jr.; McGinley, P. M.; Katz, L. E. Environ. Sci. Technol. 1992, 26, 1955-1962. (3) McGinley, P. M.; Katz, L. E.; Weber, W. J., Jr. Environ. Sci. Technol. 1993, 27, 1524-1531. (4) Young, T. M.; Weber, W. J., Jr. Environ. Sci. Technol. 1995, 29, 92-97. (5) Grathwohl, P. Environ. Sci. Technol. 1990, 24, 1687-1693. (6) Carter, M. C.; Kilduff, J. E.; Weber, W. J., Jr. Environ. Sci. Technol. 1995, 29, 1773-1780. (7) Karickhoff, S. W. In Contaminants and Sediments: Analysis, Chemistry, and Biology; Baker, R. A., Ed.; Ann Arbor Science: Ann Arbor, MI, 1980; Vol. 2, pp 193-205. (8) Miller, C. T.; Weber, W. J., Jr. J. Contam. Hydrol. 1986, 1, 243261. (9) Weber, W. J., Jr.; Miller, C. T. Water Res. 1988, 22, 457-464. (10) Miller, C. T.; Weber, W. J., Jr. Water Res. 1988, 22, 465-474. (11) Wu, S.-C.; Gschwend, P. M. Environ. Sci. Technol. 1986, 20, 717725.

(12) Crittenden J. C.; Hutzler, N. J.; Greyer, D. G.; Oravitz, J. L.; Friedman, G. Water Resour. Res. 1986, 22, 271-284. (13) Ball, W. P.; Roberts, P. V. Environ. Sci. Technol. 1991, 25, 12371249. (14) Yiacoumi S.; Tien, C. Water Resour. Res. 1994, 30, 571-580. (15) Pedit, J. A.; Miller, C. T. Environ. Sci. Technol. 1994, 28, 20942104. (16) Pedit, J. A.; Miller, C. T. Environ. Sci. Technol. 1995, 29, 17661772. (17) van Genuchten, M. T.; Davidson, J. M.; Wierenga, P. J. Soil Sci. Soc. Am. J. 1974, 38, 29-35. (18) Hinz, C.; Gaston, L. A.; Selim, H. M. Water Resour. Res. 1994, 30, 3013-3021. (19) van Genuchten, M. T.; Wierenga, P. J.; O’Connor, G. A. Soil Sci. Soc. Am. J. 1977, 41, 278-285. (20) Rao, P. S. C.; Davidson, J. M.; Jessup, R. E.; Selim, H. M. Soil Sci. Soc. Am. J. 1979, 43, 22-28. (21) van Genuchten, M. T.; Wierenga, P. J. Soil Sci. Soc. Am. J. 1976, 40, 473-480. (22) Nkedi-Kizza, P.; Biggar, J. W.; Selim, H. M.; van Genuchten, M. T.; Wierenga, P. J.; Davidson, J. M.; Nielsen, D. R. Water Resour. Res. 1984, 20, 1123-1130. (23) Hassett, J. J.; Means, J. C.; Banwart, W. L.; Wood, S. G. Sorption Properties of Sediments and Energy-Related Pollutants; EPA-600/ 3-80-041; U.S. Environmental Protection Agency: Washington, DC, 1980. (24) Means, J. C.; Wood, S. G.; Hassett, J. J.; Banwart, W. L. Environ. Sci. Technol. 1980, 14, 1524-1528. (25) Banwart, W. L.; Khan, A.; Hassett, J. J. J. Environ. Sci. Health, Part B 1980, B15, 165-179. (26) Karickhoff, S. W. Chemosphere 1981, 10, 833-846. (27) Banwart, W. L.; Hassett, J. J.; Wood, S. G.; Means, J. C. Soil Sci. 1982, 133, 42-47. (28) Karickhoff, S. W. J. Hydraul. Eng. 1984, 110, 707-735. (29) Kile, D. E.; Chiou, C. T.; Zhou, H.; Li, H.; Xu, O. Environ. Sci. Technol. 1995, 29, 1401-1406. (30) Ball, W. P.; Buehler, C.; Harmon, T. C.; Mackay, D. M.; Roberts, P. V. J. Contam. Hydrol. 1990, 5, 253-295. (31) Gregg, S. J.; Sing, K. S. W. Adsorption, Surface Area and Porosity; Academic Press: London, 1982. (32) Wauchope, R. D.; Savage, K. E.; Koskinen, W. C. Weed Sci. 1983, 31, 744-751. (33) Chemical Engineers’ Handbook, 5th ed.; McGraw-Hill Book Co.: New York, 1973. (34) Karickhoff, S. W.; Brown, D. S.; Scott, T. A. Water Res. 1979, 13, 241-248. (35) Carroll, K. M.; Harkness, M. R.; Bracco, A. A.; Balcarcel, R. R. Environ. Sci. Technol. 1994, 28, 253-258. (36) Murphy, E. M.; Zachara, J. M.; Smith, S. C. Environ. Sci. Technol. 1990, 24, 1507-1516. (37) Murphy, E. M.; Zachara, J. M.; Smith, S. C.; Philips, J. L.; Wiestma, T. W. Environ. Sci. Technol. 1994, 28, 1291-1299. (38) Gu, B.; Schmitt, J.; Chen, Z.; Liang, L.; McCarthy, J. F. Environ. Sci. Technol. 1994, 28, 38-46. (39) Engebretson, R. R.; von Wandruszka, R. Environ. Sci. Technol. 1994, 28, 1934-1941. (40) Wershaw, R. L. Environ. Sci. Technol. 1993, 27, 814-816. (41) Wershaw, R. L. In Organic Substances and Sediments in Water. Volume 1. Humics and Soils; Baker, R. A., Ed.; Lewis Publishers: Chelsea, MI, 1991; pp 23-34. (42) Wershaw, R. L. J. Contam. Hydrol. 1986, 1, 29-45. (43) Durand, B. Kerogen: Insoluble Organic Matter from Sedimentary Rocks; Technip: Paris, 1980. (44) Tissot, B. P.; Welte, D. H. Petroleum Formation and Occurrence; Springer-Verlag: New York, 1984. (45) Engel, M. H.; Macko, S. A. Organic Geochemistry: Principles and Applications; Plenum Press: New York, 1993. (46) Schnitzer, M. In Soil Organic Matter; Schnitzer, M., Khan, S. U., Eds.; Elsevier: Amsterdam, 1978. (47) Schnitzer, M.; Kodama, H.; Ripmeester, J. A. Soil Sci. Soc. Am. J. 1991, 55, 745-750. (48) Grathwohl, P.; Reinhard, M. Environ. Sci. Technol. 1993, 27, 23602366. (49) Connaughton, D. F.; Stedinger, J. R.; Lion, L. W.; Shuler, M. L. Environ. Sci. Technol. 1993, 27, 2397-2403. (50) Crank, J. The Mathematics of Diffusion, 2nd ed.; Oxford University Press: Oxford, U.K., 1975. (51) Crank, J.; Park, G. S. Diffusion in Polymers; Academic Press: London and New York, 1968. (52) Brusseau, M. L.; Jessup, R. E.; Rao. P. S. Environ. Sci. Technol. 1991, 25 134-142. (53) Pignatello, J. J. In Reactions and Movements of Organic Chemicals in Soils; Sawhney, B. L., Brown, K., Eds.; Soil Science Society of


9

887

(54) (55) (56) (57) (58)

888

America, Inc. and American Society of Agronomy, Inc.: Madison, WI, 1989; Chapter 3. Schwarzenbach, R. P.; Gschwend, P. M.; Imboden, D. M. Environmental Organic Chemistry; John Wiley & Sons, Inc.: New York, 1993. Schwarzenbach, R. P.; Westall, J. Environ. Sci. Technol. 1981, 15, 1360-1367. Schlautman, M. A.; Morgan, J. J. Environ. Sci. Technol. 1994, 28, 2184-2190. McCarthy, J. F.; Jimenez, B. D. Environ. Sci. Technol. 1985, 19, 1072-1076. Gauthier, T. D.; Seitz, W. R.; Grant, C. L. Environ. Sci. Technol.

9


1987, 21, 243-248. (59) Backhus, D. A.; Gschwend, P. M. Environ. Sci. Technol. 1990, 24, 1214-1223. (60) Schlautman, M. A.; Morgan, J. J. Environ. Sci. Technol. 1993, 27, 961-969.

Received for review May 15, 1995. Revised manuscript received October 23, 1995. Accepted November 3, 1995.X ES950329Y X

Abstract published in Advance ACS Abstracts, January 15, 1996.

A Distributed Reactivity Model for Sorption by Soils and Sediments. 4

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