Environ. Sci. Technol. 2009, 43, 1930–1934
Herbicide Sorption by Immersed Soils: Stoichiometry and the Law of Mass Action in Support of Predictive Kinetics DONALD S. GAMBLE* Department of Chemistry, Saint Mary’s University, Halifax, Nova Scotia B3H 3C3
Received September 8, 2008. Revised manuscript received December 17, 2008. Accepted January 20, 2009.
The stoichiometry of labile herbicide sorption on immersed soils has been determined for a few herbicides and a number of soils (Gamble, D. S.; Khan, S. U. Atrazine in organic soil: Chemical speciation during heterogeneous catalysis. J. Agric. Food Chem. 1990, 38, 297-308; Gamble, D. S.; Ismaily, L. A. Atrazine in mineral soil: The analytical chemistry of speciation. Can. J. Chem. 1992, 70, 1590-1596; Gamble, D. S.; Khan, S. U. Atrazine in mineral soil: Chemical species and catalysed hydrolysis. Can. J. Chem. 1992, 70, 1597-1603; Gilchrist, G. F. R.; Gamble, D. S.; Khan, S. U. Atrazine interactions with clay minerals: Kinetics and equilibria of sorption. J. Agric. Food Chem. 1993, 41, 1748-1755; Gamble, D. S. Physical chemistry parameters that control pesticide persistence and leaching in watershed soils. Final report submitted to the Great Lakes Water Quality Program Committee, Guelph, Ontario, June, 1994; Li, J.; Langford, C. H.; Gamble, D. S. Atrazine sorption by a mineral soil: Processes of labile and nonlabile uptake. J. Agric. Food Chem. 1996, 44, 3672-3679; Li, J.; Langford, C. H.; Gamble, D. S. Atrazine sorption by a mineral soil: The effects of size fractions and temperature. J. Agric. Food Chem. 1996, 44, 3680-3684; Gamble, D. S. Pesticide-soil research for the behaviour of chlorothalonil and its metabolite SD-3701 in soil. Final report submitted to Ricerca Inc., Sept. 15, 1998; Gamble, D. S. Atrazine sorption kinetics in a characterized soil: Predictive calculations. Environ. Sci. Technol. 2008, 42, 1537-1541). This was done by using equilibrium titrations for the measurement of labile sorption capacities θC. The titrations were made possible by resolving total sorption into its labile and unrecovered fractions. But equilibrium was not also necessary for unrecovered fractions. The site saturation at titration plateaus defined θC. The first purpose of determining the stoichiometry is to permit the use of second-order kinetics instead of unpredictive pseudo-first-order kinetics for sorption modeling. Another purpose is to replace empirical distribution coefficients such as KD with the law of mass action for describing equilibria as limiting states. Temperature trends and a comparison with EGME vapor deposition data from the literature indicate a control of herbicide sorption by sorbed water. A preliminary examination of limited data from different sources suggests that future research should investigate some additional correlations. θC and equilibrium functions might both be
* E-mail:
[email protected]. 1930
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 6, 2009
influenced by soil organic matter carboxyls and carboxylate anions, as well as inorganic materials. Some disadvantages of KD are noted.
Introduction The behavior of pesticides in soils and sediments is known to be dynamic. Equilibria will exist infrequently if at all (9, 10). But equilibria and irreversible states define the limiting conditions that can be kinetically approached. The authors of two critical reviews have independently concluded that empirical models and the empirical equilibrium parameters KD, KOC, and KOW have practical disadvantages for predictive calculations (11, 12). Starting with sorption as the basis for more complete mechanism studies, it has been shown that chemical stoichiometry and the law of mass action are possible for the quantitative predictions of pesticides in soils (9). Two requirements for this have been demonstrated. First, the titrations of immersed soils with hydrophobic pesticides have revealed labile sorption capacities, θC (4-8). This established the chemical stoichiometry. Second, equilibrium as a limiting condition can be described by adapting the law of mass action to the mixture of sorption sites found in a natural soil. Versions of the theory already exist for cations in soils and soil components (13, 14). A very large amount of literature exists for pesticide sorption equilibria in soils. Two recent reviews are important because they present critical examinations of the concepts and methods currently in universal use. In both of these reviews and in all of the publications that they cite, it has generally been assumed that only equilibrium parameters would be used for operational and regulatory purposes. It was also generally assumed that the empirical equations and parameters were the only available options. A IUPAC project, no. 640/43/97, produced a critical review of pesticide sorption parameters. The concepts, methods, and applications were examined by three IUPAC commissions in 2002 (15). They were “Agrochemicals and the Environment”, “Fundamental Environmental Chemistry”, and “Soil and Water Chemistry”. It was noted that the distribution coefficient KD and the related KOC are in worldwide use for data collection and regulatory work. The reviewers pointed out that while KOC is assumed to be valid for all soils, this is known to be an approximation. The limitations of these parameters were reported, together with some recommendations for their use in fate and transport models. Sitea published the other comprehensive review in 2001 (15). He presented a critical examination of 681 articles on the soil sorption properties of hydrophobic organic molecules. Tabulated log KD and log KOC were correlated in some cases with organic C, soil pH, and clay content. Those reviews are only a representative sampling of the large pesticide-soil sorption literature. Several other publications confirm the dependence of sorption on both organic matter and clays (16-20). Since the reviews, evidence of hydrophobicity and cation exchange effects on sorption has been published (21-24). This literature defines the conventional state of the art as of 2001. The experimental methods reported in these reviews do not include any method for resolving total sorption into labile sorbed and bound residue fractions. This is in spite of the fact that bound residues have been reported by several authors since the 1966 article by Hamaker et al. (25-41). The large accumulations of empirical KD and KOC data therefore seem to represent ambiguous combinations of labile and nonlabile sorption. Also none of the methods discussed include any that measure labile sorption ca10.1021/es8025177 CCC: $40.75
2009 American Chemical Society
Published on Web 02/11/2009
pacities, θC. That is, no methods are described for determining the (mol/g) of total sites on immersed soils available for labile sorption of hydrophobic molecules. These are the stoichiometric sorption sites. Instead, the term “capacity” is only used to refer to the Freundlich isotherm parameter. This empirical parameter has no known relationship to the total number of labile sorption sites, as measured by site saturation. The conventional state of the art as of 2002 precluded the use of chemical stoichiometry for the quantitative prediction of dynamic conditions and their equilibrium limits. Without the resolution of total sorption into labile and nonlabile fractions and without experimental data for the numbers of filled and unoccupied labile sorption sites, the law of mass action cannot be adapted to natural soil mixtures. The Polanyi-Manes model has been used for describing equilibrium sorption from solution (26). It gives the energy difference between two processes: (a) the energy of solute escaped into saturated solution from the surface of its own solid; (b) the energy of solute escaped into subsaturated solution from the surfaces of one or more other solids. Future research could consider whether or not to use it for sorption onto soils. Like KD, it does not use stoichiometry with which to account for the law of mass action. It has therefore not been used in this work. This report has two objectives in support of the previously reported kinetics of pesticide sorption onto immersed soils (9). The first is to demonstrate that the stoichiometry of labile sorption can be determined, so that second-order kinetics can be used instead of unpredictive pseudo-first-order kinetics (9). Some labile sorption capacities θC required for this purpose have been tabulated (Supporting Information). The other objective is to show that equilibria as limiting states can be more correctly described with the law of mass action instead of the usual distribution coefficients KD. Some weighted average equilibrium functions have been tabulated for the mixtures of sorption sites found in soils (Supporting Information). These concepts were evaluated using a Plainfield soil collected from Southern Ontario with the herbicides atrazine and metolachlor.
Theory Equilibrium titrations of immersed solids with hydrophobic molecules have revealed limits to the amounts that could be reversibly sorbed. To distinguish this type of physical sorption limit from sorption that is reversible either slowly or not at all, it is referred as the labile sorption capacity, θC (mol/g) (13). Sorbed water competes with the hydrophobic molecules for labile sorption sites. The equilibrium between sorbed water and the sorbed hydrophobic molecules causes the apparent labile sorption capacity, θC (mol/g). Titrations of hydrophobic molecules onto the available sites reveal θC as saturation limits at titration curve plateaus. Chemical stoichiometry is established by recognizing that the sorption sites occupied by hydrophobic molecules and those occupied by water are reactants. This permits the calculation of weighted average equilibrium functions. For soils and sediments with mixtures of surface sorption sites, the presaturation sections of the titration j 1. curves yield a law of mass action equilibrium function K This is a weighted average equilibrium function that is measured experimentally for the whole mixture of sorption sites (13, 14). Sorption equilibrium is governed by the law of mass action, as described by eq 3. The theory previously developed for cations and humic materials has been adapted here for hydrophobic molecules in immersed soils (13, 14). In terms of the concentrations of dissolved
chemical MS and water filled sorption sites MW as reactants, the equilibrium can be described by eq 1. MS+MW a M1+ML
K ′ ≈ (M1ML)/(MSMW)
(1)
The reaction products are the sorption sites M1 occupied by the chemical, and by water ML in the bulk solution. But for the case of solids immersed in dilute solution, ML is approximately constant with aW ≈ 1. Also, eqs 2a and 2b account for the ratio of solid to solution. M1 ) (w/v)θ1
(2a)
MW ) (w/v)θW
(2b)
θ1 and θW are the (mol/g) of the chemical and of water on the labile sorption sites. (w/v) is the (g of solids/L of solution). Equation 1 is revised to give eq 3. j 1 ) θ1/MSθW K
(3)
A soil or sediment has a collection of nonidentical labile sorption sites. For the ith small portion of them and its equilibrium, they are θiC ) θi1 + θiW
(4)
Ki ) θi1/MSθiW
(5)
The sum of sites filled by the chemical is given by eq 6. n
θ1 )
∑θ
(6)
i1
i)1
j 1 as a weighted average equilibrium These relationships give K function in eq 7. n
j 1 ) (1/θw) K
∑Kθ
(7)
i iW
i)1
Assuming that the sizes of the Ki values and θiW intervals are sufficiently close to avoid step functions, the summation may be modeled as an integral (15). And using an experimental value for θC for conversion to a mole fraction form, the weighted average is expressed as eq 8. j 1 ) (1/XW) K
∫
XW
0
K dXW
(8)
This allows the differential equilibrium function K1 to be calculated with eq 9 and used for the estimation of ∆G° (13). j 1 cannot be properly used for the calculation because K chemical potentials are defined only for individual reactants rather than mixtures (13). j 1XW)/dXW K1 ) d(K
(9)
Results And Discussion The Plainfield soil was collected from southwestern Ontario. The analyses by X-ray diffraction and EGME are tabulated in Table S1. EGME measurements are conventionally interpreted as surface area, (m2/g). But because BET and EGME measurements generally give different numerical values and because (m2/g) does not support chemical stoichiometry, the EGME measurements have been expressed as (mol/g) for comparison with θC (mol/g). Figure 1 compares the structures of atrazine and metolachlor. Metolachlor is slightly more soluble than atrazine, which indicates that it is slightly less hydrophobic. Labile sorption properties would be expected to be different. Figure 2 shows an example of the titrations of Plainfield soil with atrazine. Especially at the beginning of the titration, it was important for the total sorption to be resolved into its labile VOL. 43, NO. 6, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
1931
TABLE 1. Labile Sorption Capacities of Immersed Plainfield Soil
FIGURE 1. Atrazine and metolachlor.
chemical
temp. (°C)
labile sorption capacity θC(mol/g)
std dev (mol/g)
no. of data
atrazine atrazine atrazine metolachlor
10 25 35 25
2.23E-08 8.20E-09 5.30E-09 2.37E-08
1.0E-09 8.6E-10 4.4E-10 3.5E-10
24 8 12 13
sorption onto dry surfaces. A temperature increase would of course tend to desorb some organic chemical from a dry surface. But the experiments indicate that for solvated surfaces, a temperature increase desorbs some of the water. This has allowed an increased sorption of the hydrophobic organic molecules. The temperature dependent sorption of water explains why atrazine and metolachlor show similar effects. Finally, the comparison of the metolachlor and atrazine curves at 25 °C shows that in this case the more soluble compound was also more sorbed onto the labile sorption sites. This raises questions about the competitive interactions of hydrophobic molecules with bulk water and solvated surfaces. Octanol-water partition coefficients based on alcoholic OH groups cannot account well for the pH dependent protonation of humic carboxyl groups. The commonly used logarithmic plots obscure any scatter in experimental data. A 10-fold error makes a difference of only 1 log unit. Table 1 has the values calculated for Figures 2 and 3. In some cases solubility limits make the titration of labile sorption sites more difficult. j 1θ1 + K j 1θC KD ) K
FIGURE 2. Titration of Plainfield soil labile sorption sites by aliquot additions, 10 °C.
FIGURE 3. Titrations of Plainfield soil with atrazine and metolachlor. Curve plateaus show the saturation of labile sorption sites, giving the labile sorption capacities, θC. and nonlabile fractions. While labile sorption equilibrium was required for each aliquot addition, the intrapartricle sorption by diffusion was not necessarily at equilibrium. Table S2 has the experimental data for the 10 °C titration curve in Figure 2. Table S3 has the data for the similar 25 °C titration of metolachlor. Three facts are revealed by Figure 3. First, four independent experiments confirm the existence of labile sorption capacities for two herbicides by an immersed soil. Second, the temperature trend for the atrazine curves should be noted. This trend is opposite to that which would be expected for 1932
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 6, 2009
(10)
With a suitable data set, θC can be estimated with eq 10 (15). But large standard deviations or nonlinear data can give unreliable estimates. The atrazine estimates with eq 10 are too high in the present examples, because of nonlinearity. A small but significant database of θC measurements has been assembled in Table S4. Published and unpublished data are included, as indicated by the references. Some trends suggested by the limited data in Figures S1-S4 are consistent with the sorption research of other authors, but experimental confirmation is needed. Significant effects of organic matter, clays, and pH have been reported by other authors, for sorption parameters other than θC (15-21). As previously reported (1, 5, 13, 14), protonated carboxyl groups in undissolved soil organic matter contribute to the labile sorption of hydrophobic molecules while carboxylate anions do not. The solution concentration of H+, as indicated by pH, does not directly influence θC. Instead it controls the carboxyl protonation which directly influences θC. j 1 in Figure The weighted average equilibrium function K 4 has been calculated with equilibrium measurements below j j1 unexpectedly shows no temperature trend. site saturation. K The anticipated temperature effect is found instead in θC, in Figure 3. Figure 4 also shows another unexpected trend. For j 1 would decrease as the mole sorption onto a dry surface, K fraction of sites occupied by the herbicides increased. This is because the most strongly sorbing sites would be occupied first. But for herbicide sorption onto immersed surfaces, it j 1 decreased with was instead found experimentally that K decreasing mole fraction of sites occupied by the herbicides. That is, it decreased with increasing mole fraction of sites occupied by water. The data for Figure 4 are tabulated in Table S5, for future use. The reason for the unexpected temperature trends is that labile sorption is controlled by water which competes with the hydrophobic molecules for sorption sites. It is this
FIGURE 4. Weighted average equilibrium function K¯j1, increasing with decreasing water occupation of labile sorption sites. The three atrazine data sets were fitted together, and metolachlor data were plotted separately. Relative standard deviation ) 10%.
FIGURE 5. 9, Weighted average equilibrium function K¯j1, relative standard deviation ) 10%. b, Differential equilibrium function K1, relative standard deviation ) 20%. competition between water and hydrophobic molecules that causes the apparent labile sorption capacity θC to exist. There is an analogue of this in ion exchange chemistry. The apparent cation exchange capacity of carboxylic ion exchangers is controlled by the loading of the carboxyls with competing protons (41). This suggests that θC might increase with decreasing moisture content. Although this has practical implications for pesticide fate and transport models, it has never been experimentally investigated. Molecular modeling for this might be more difficult than the work done by Shevchenko and Bailey (42) for dry surfaces. j 1 would be appropriate for hydrology engineering K calculations. It would not, however, be the correct equilibrium function for the calculation of thermodynamic functions or for molecular level interpretations. The differential equilibrium function K1 defined by eq 9 is more correct for those purposes (13, 14). The extent to which K1 differs from the weighted average equilibrium function in a particular case depends on the nature of the mixture of reactants. For the example of atrazine in Plainfield soil, Figure 5 shows that at least 30% of the strongest sorbing sites are different from each other. The rest of the labile sorption sites appear to be j 1 reduces approximately energetically rather similar, so that K to K1. But this is not always the case. Weak sorption might have obscured any differences. The ∆G ° values estimated from K1 are approximately 30-50 kJ/mol, which are within the range of physical interactions such as hydrogen bonding. Some of the difficulties of KD for nonequilibrium and low coverage of sorption sites have been reported (9). But even if sorption equilibrium were to exist, there would still be problems with the use of constant KD. The first is that it does not account for the law of mass action. This neglect of θC can cause significant errors because its values range over orders of magnitude, as seen in Table S3. Figure 6 gives typical examples of another problem. The equilibrium values from
FIGURE 6. Distribution coefficients KD. The effect of temperature on the equilibrium values. σ: b, 8.1E-04; 9, 4.2E-04; 2, 3.0E-04; X in square, 5.8E-04. four independent experiments are not constant. The reason is that a natural soil is a mixture of materials. A range of sorption energies therefore exists, and the most strongly sorbing sites are occupied first as solution concentration increases. If the KD in Figure 6 is made to look approximately constant within its standard deviation σ by confining the measurements to a narrow window, then the numerical value will be different for a different narrow window. The resulting errors in predictive calculations would be calculation artifacts rather than random measurement errors. The labile sorption capacities, θC, can be measured for some combinations of pesticides and soils. For natural soils and sediments, the mixtures of labile sorption sites experimentally produce equilibrium functions that are weighted average functions. When fitted to the mole fraction of occupied sorption sites, these weighed average functions generally yield distribution curves. The concepts and methods for chemical stoichiometry and equilibria described here further support the previously reported kinetics model (9). It is proposed that for other kinetics models, chemical stoichiometry and the law of mass action should replace, when possible, less predictive empirical parameters such as Freundlich isotherms and distribution coefficients. Future research should also expand and correct the databases of labile sorption capacities, θC, and law of mass action equilibrium functions (9). The conclusions apply directly to saturated soil conditions and aquatic sediments. In the unsaturated soil zone, they might also apply to conditions in which aqueous solutions were in contact with totally wet soil surfaces.
Supporting Information Available Tables S1-S5 and Figures S1-S4. This material is available free of charge via the Internet at http://pubs.acs.org.
Acknowledgments The Research Branch of Agriculture and Agri-Food Canada, the former Ricerca Inc. of Painesville, OH, the Department of Chemistry and Biochemistry of Concordia University, the Department of Chemistry of the University of Calgary, and the Department of Chemistry of Saint Mary’s University have each supported various parts of this research.
Literature Cited (1) Gamble, D. S.; Khan, S. U. Atrazine in organic soil: Chemical speciation during heterogeneous catalysis. J. Agric. Food Chem. 1990, 38, 297–308. (2) Gamble, D. S.; Ismaily, L. A. Atrazine in mineral soil: The analytical chemistry of speciation. Can. J. Chem. 1992, 70, 1590– 1596. (3) Gamble, D. S.; Khan, S. U. Atrazine in mineral soil: Chemical species and catalysed hydrolysis. Can. J. Chem. 1992, 70, 1597– 1603. VOL. 43, NO. 6, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
9
1933
(4) Gilchrist, G. F. R.; Gamble, D. S.; Kodama, H.; Khan, S. U. Atrazine interactions with clay minerals: Kinetics and equilibria of sorption. J. Agric. Food Chem. 1993, 41, 1748–1755. (5) Gamble, D. S. Physical chemistry parameters that control pesticide persistence and leaching in watershed soils. Final report submitted to the Great Lakes Water Quality Program Committee, Guelph, Ontario, June 27, 1994. (6) Li, J.; Langford, C. H.; Gamble, D. S. Atrazine sorption by a mineral soil: Processes of labile and nonlabile uptake. J. Agric. Food Chem. 1996, 44, 3672–3679. (7) Li, J.; Langford, C. H.; Gamble, D. S. Atrazine sorption by a mineral soil: The effects of size fractions and temperature. J. Agric. Food Chem. 1996, 44, 3680–3684. (8) Gamble, D. S. Pesticide-soil research for the behaviour of chlorothalonil and its metabolite SD-3701 in soil. Final report submitted to Ricerca Inc., Sept. 15, 1998. (9) Gamble, D. S. Atrazine sorption kinetics in a characterized soil: Predictive calculations. Environ. Sci. Technol. 2008, 42, 1537– 1541. (10) Karickhoff, S. W.; Morris, K. R. Sorption dynamics of hydrophonic pollutants in sediment suspension. Environ. Toxicol. Chem. 1985, 4, 469–479. (11) Baskaran, S.; Bolan, N. S.; Rahman, A.; Tillman, R. W. Pesticide sorption by allophanic and non-allophanic soils of New Zealand. New Zealand J. Agric. Res. 1996, 39, 297–310. (12) Yeh, S.; Linders, J. B. H. J.; Kloskowski, R.; Tanaka, K.; Rubin, B.; Katayama, A.; Ko¨rdel, W.; Gerstl, Z.; Lane, M.; Unsworth, J. B. Pesticide soil sorption parameters: theory, measurement, uses, limitations and reliability. IUPAC project, no. 640/43/97. Soc. Chem. Ind. 2002, 58 (5), 419–445. (13) Gamble, D. S.; Langford, C. H.; Webster, G. R. B. Interactions of pesticides and metal ions with soils: Unifying concepts. Rev. Environ. Contam. Toxicol. 1994, 135, 63–91. (14) Gamble, D. S.; Langford, C. H.; Bruccoleri A. G. Chemical stoichiometry and molecular level mechanisms as support for future predictive engineering. Chapter 6 in Use of humic substances to remediate polluted environments: From theory to practice. NATO Science Series; Perminova, I. V., Herkorn, N., Baveye, P., Eds.; Kluwer Academic Publisher: Dordrecht, 2005. (15) Sitea, A. D. Factors affecting sorption of organic compounds in natural sorbent water systems and sorption coefficients for selected pollutants. A Review. J. Phys. Chem. Ref. Data. 2001, 30 (1), 187–439. (16) Bengtsson, S.; Berglo¨f, T.; Kylin, H. Near infrared reflectance spectroscopy as a tool to predict pesticide sorption in soil. Bull. Environ. Contam. Toxicol. 2007, 78 (5), 295–298. (17) Li, H.; Sheng, G.; Teppen, B. J.; Johnston, C. T.; Boyd, S. A. Sorption and desorption of pesticides by clay minerals and humic acid-clay complexes. Soil Sci. Soc. Am. J. 2003, 67, 122– 131. (18) Weber, J. B.; Wilkerson, G. G.; Reinhardt, C. F. Calculating pesticide sorption coefficients (Kd) using selected soil properties. Chemosphere 2004, 55 (2), 157–166. (19) Yu, Y.; Zhou, Q. X. Adsorption characteristics of pesticides methamidophos and glyphosate by two soils. Chemosphere 2005, 58, 811–806. (20) Zbytniewski, R.; Buszewski, B. Sorption of pesticides in soil and compost. Pol. J. Environ. Stud. 2002, 11 (2), 179–184. (21) Spadotto, C. A.; Hornsby, A. G. Soil sorption of acidic pesticides: Modeling pH effects. J. Environ. Qual. 2003, 32, 949–956. (22) de Oliveira, M. F.; Prates, H. T.; Santanna, D. P.; de Oliveira Ju ´ nior, R. S. Imazaquin sorption in surface and subsurface soil samples. Pesq. agropec. bras., Brası´lia 2006, 41 (3), 461–468.
1934
9
ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 6, 2009
(23) Li, H.; Teppen, B. J.; Laird, D. A.; Johnston, C. T.; Boyd, S. A. Geochemical modulation of pesticide sorption on smectite clay. Environ. Sci. Technol. 2004, 38 (20), 5393–5399. (24) Sa´nchez, L.; Romero, E.1; Sa´nchez-Rasero, F.; Dios, P. A. Enhanced soil sorption of methidathion. Pest Management Sci. 2003, 59 (8), 857–864. (25) Hamaker, J. W.; Goring, C. A. L.; Youngson, C. R. Degradation of atrazine and related S-triazines. Organic pesticides in the environment; American Chemical Society: Washington, DC, 1996; p 23. (26) Manes, M. Activated carbon adsorption fundamentals. In Encyclopedia of environmental analysis and remediation; Myers, R. A., Ed.; Wiley: New York, 1998; pp 26-68. (27) Hamaker, J. W.; Thompson, J. M. Adsorption in organic chemicals in the soil environments; Goring, C. A. L., Hamaker, J. W., Eds.; Dekker: New York, 1972; Vol. 1, pp 49-143. (28) Kearney, P. C. Summary of soil bound residue sessions. ACS Symp. Ser. 1976, 29, 378–382. (29) Wu, S.-C.; Gschwend, P. M. Sorption kinetics of hydrophobic organic compounds to natural sediments and soils. Environ. Sci. Technol. 1986, 20, 717–725. (30) Steinberg, S. M.; Pignatello, J. J.; Sawhney, B. L. Persistence of 1,2-dibromoethane in soils entrapment in intraparticle microspace. Environ. Sci. Technol. 1987, 21, 1201–1208. (31) Weber, W. J., Jr.; McGinley, P. M.; Katz, L. E. Sorption phenomena in subsurface systems: Concepts, models and effects on contaminant fate and transport. Water Res. 1987, 25, 499–528. (32) Clay, S. A.; Allmaras, R. R.; Koskinen, W. C.; Wise, D. L. Desorption of atrazine and cyanazine from soil. J. Environ. Qual. 1988, 17 (4), 719–723. (33) Brusseau, M. L.; Rao, P. C. S. Degradation of atrazine and related S-triazines. Crit. Rev. Environ. Control 1989, 19, 1–14. (34) Ball, W. P.; Roberts, P. V. Long-term sorption of halogenated organic chemicals by aquifer material. 1. Equilibrium. Environ. Sci. Technol. 1991a, 25, 1223–1237. (35) Ball, W. P.; Roberts, P. V. Long-term sorption of halogenated organic chemicals by aquifer material. 2. Intraparticle diffusion. Environ. Sci. Technol. 1991b, 25, 1237–1249. (36) Weber, W. J.; McGinley, P. M.; Katz, L. E. A distributed reactivity model for sorption by soils and sediments. Conceptual basis and equilibrium assessments. Environ. Sci. Technol. 1992, 26, 1955–1962. (37) Li, J. Equilibrium and kinetics studies of atrazine and lindane uptake by soils and soil components. Doctoral Thesis, Department of Chemistry and Biochemistry, Concordia University, Montreal, 1993. (38) Kan, A. T.; Fu, G.; Thompson, M. B. Adsorption/desorption hysteresis in organic pollutant and soil/sediment interaction. Environ. Sci. Technol. 1994, 28, 859–867. (39) Pignatello, J. J.; Xing, B. Mechanisms of slow sorption to natural particles. Environ. Sci. Technol. 1996, 30, 1–11. (40) Hu, M. Q.; Brusseau, M. L. Coupled effects on nonlinear, ratelimited sorption and biodegradation on transport of 2, 4-dichlorophenoxyacetic acid in soil. Environ. Toxicol. Chem. 1998, 17, 163–168. (41) Helfferich, F. Ion Exchange; McGraw-Hill Book Company, Inc.: New York, 1962; pp 90-94. (42) Shevchenko, S. M.; Bailey, G. W. Non-bonded organo-mineral interactions and sorption of organic compounds on soil surfaces: A model approach. J. Mol. Struct. (Theochem) 1998, 422, 259–270.
ES8025177
