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Masters Theses 1911 - February 2014
1999
Sorption/desorption of organic compounds by soil organic matter /
Guoshu Yuan University of Massachusetts Amherst
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SORPTION/DESORPTION OF ORGANIC COMPOUNDS BY SOIL ORGANIC MATTER
A Thesis Presented
by
GUOSHU YUAN
Submitted to the Graduate School of the University of Massachusetts Amherst in partial fulfillment of the requirements for the degree of
MASTER OF SCIENCE
September 1999
Plant and Soil Sciences SORPTION/DESORPTION OF ORGANIC COMPOUNDS BY SOIL ORGANIC MATTER
A Thesis Presented
by
GUOSHU YUAN
Approved as to style and content by: ACKNOWXEDGMENTS
I would like to thank my research advisor. Dr. Baoshan Xing for his guidance and support throughout the course of my study in Plant and Soil Sciences Department. His expertise in the field of soil science helped me to choose a challenging topic to work with. This working experience will benefit me in my future career.
1 wish to thank Dr. Stephen Simkins for his insight and technical help for my research. I convey my gratitude to Dr. Prasanta Bhowmik and Dr. Sharon Long for their invaluable advice.
Friendship from Zhaowei, Kun, Jingdong, Sowmya, Amrith and Sanjay is greatly appreciated.
Finally, 1 want to thank my parents and brothers for their love and encouragement.
Ill TABLE OF CONTENTS
I^agc
ACKNOWLEDGMENTS.iii
LIST OF TABLES.vi
LIST OF FIGURES.vii
CHAPTER
L GENERAL INTRODUCTION.1
Literature Review.3 Objectives.10 References.11
2. EFFECTS OF METAL CATIONS ON SORPTION/DESORPTION OF ORGANIC COMPOUNDS BY HUMIC ACIDS.14
Abstract.14 Introduction.15 Materials and Methods.16 Results and Discussions.21 References.35
3. SITE-ENERGY DISTRIBUTION ANALYSIS OF ORGANIC CHEMICAL SORPTION BY SOIL ORGANIC MATTER.37
Abstract.37 Introduction.37 Theoretical.40 Experimental.44 Results and Discussions.47 References.54
IV 4. SYNTHESIS 57
Summary.57 Significance in the Field.58 Future Research.59
APPENDICES
A. '^C NMR SPECTRA OF HA.60 B. '^C NMR CHARACTERIZATION OF HA.61
BIBLIOGRAPHY.62
V LIST OF TABLES
Table Page
2.1 Properties of model compounds.17
2.2 Sorption parameters and calculated Kqc values for Amherst HA.22
2.3 Desorption parameters for chemicals on Amherst HA.25
2.4 Hysteresis indices for chemicals in Amherst HA.31
2.5 Hysteresis Indices for naphthalene in Chelsea HA.34
3.1 Composition of sorbents.45
3.2 Freundlich isotherm parameters.48
VI LIST OF FIGURES
Figure Page
2.1 Successive desorption of naphthalene by Amherst HA.26
2.2 Successive desorption of naphthalene by Amherst Ca-HA.26
2.3 Successive desorption of naphthalene by Amherst Al-HA.27
2.4 Successive desorption of phenanthrene by Amherst HA.27
2.5 Successive desorption of phenanthrene by Amherst Ca-HA.28
2.6 Successive desorption of phenanthrene by Amherst Al-HA.28
2.7 Successive desorption of a-naphthol by Amherst HA.29
2.8 Successive desorption of a-naphthol by Amherst Ca-HA.29
2.9 Successive desorption of a-naphthol by Amherst Al-HA.30
2.10 Correlation of sorption nonlinearity (Ns) with hysteresis index {NJNs) at the highest initial concentrations for the three chemicals.33
2.11 Correlation of sorption nonlinearity (Ns) with hysteresis index (NJNs) at the intermediate initial concentrations for the three chemicals.33
2.12 Correlation of sorption nonlinearity (Ng) with hysteresis index {Nd/Ns) at the lowest initial concentrations for the three chemicals.34
3.1 Sorption of naphthalene by different fractions of Pahokee peat soil (PPS).49
3.2 Dependence of sorption energy on solute loading for different experimental systems.50
3.3 Approximate site-energy distributions derived from Freundlich isotherm parameters.52
Vll CHAPTER 1
GENERAL INTRODUCTION
Many soils and subsurface systems are contaminated with organic compounds as results of improper waste disposals, accidental spills, and leakage from underground storage tanks. These compounds, including pesticides, petrochemicals, industrial wastes and so on, can be very toxic to plants, animals and human beings. Therefore, it is essential to understand the transport and fate of these compounds in the environment, which will allow us to clean-up the contaminated-sites successfully.
Among the many processes influencing the transport of organic compounds, such as advection, dispersion, diffusion, and sorption/desorption, the latter is most important, because it governs the distribution of the organic compounds between the bio-accessible, mobile aqueous phase and the bio-inaccessible, immobile solid phase. Studies have shown that organic compounds preferentially sorb to soil organic matter (SOM) over the mineral phase of soil in aqueous systems due to the strong competition of water molecules for mineral surfaces (Chiou, 1989; Karickhoff, 1984). Sorption is a general term for uptake of a solute by SOM without implication of specific mechanisms.
Additional terms are also used in sorption studies. Adsorption refers to condensation of solutes on surfaces or interior pores of SOM; partition refers to uptake of a solute into the three-dimensional SOM network (Chiou, 1989).
Although sorption/desorption phenomenon has been extensively studied for many years, no consensus on sorption mechanisms has been reached to date. Sorption/desorption
1 was often simplified by assuming rapid equilibrium, isotherm linearity and desorption reversibility (Brusseau and Rao, 1989). However, data deviating from the simple assumptions have been reported, including slow kinetics of sorption and/or desorption
(Steinberg et al., 1987; Pavlostathis and Mathavan, 1992; Xing and Pignatello, 1996),
isotherm nonlinearity (Spurlock et al., 1995; Weber and Huang, 1996; Xing and Pignatello,
1997), and sorption-desorption hysteresis, i.e., the desorption isotherm not being coincident
with the sorption isotherm (Di Toro and Horzempa, 1982; Kan et al., 1994; Huang and
Weber, 1997).
Reliable estimates of sorption parameters are essential for the transport modeling.
Using nonlinear fitting sorption parameters in the transport equation will lead to
significant variations in predicting the movement of contaminants in the environment
compared to the use of the linear fitting (Brusseau and Rao, 1989).
Sorption mechanisms are mainly determined by the interactions between the
sorbing chemicals and SOM. Ionic or polar organic compounds may interact with SOM
by specific forces such as electrostatic force, covalent bond, charge transfer, or hydrogen
bond. Nonpolar compounds interact with SOM mainly by weak van der Waals forces.
Sorption isotherms were widely used to prove or reject hypothesized sorption
mechanisms (Chiou et al., 1979, 1983; Xing and Pignatello, 1997). However, one set of
data often fitted well with several equations (linear as well as nonlinear). Therefore, other
evidence such as desorption data needs to be explored to support a given hypothesis.
The present research was designed to obtain a good understanding of the
interactions between organic compounds and SOM through a series of
sorption/desorption studies. In Chapter 2, sorption/desorption of selected model
2 compounds including naphthalene, phenanthrene and a-naphthol by cation-saturated humic acids was investigated using batch equilibration method. Effects of saturating cations as well as effects of molecular size and polarity of solute on sorption/desorption were explored. In Chapter 3, a site-energy distribution analysis (Carter et al., 1995) was performed and sorption phenomenon was interpreted from a theoretical viewpoint. The
results here will advance the knowledge of sorption mechanisms and clear some
uncertainties on this subject.
Literature Review
Sorption/desorption is an underlying process affecting the transport and fate of
organic contaminants and pesticides in the environment. This field of research has gained
much recognition recently because of the public concern over the likelihood of
environmental contamination by a wide variety of organic compounds. However, no
general agreement on sorption mechanisms have been reached so far. In modeling
contaminant transport in the environment, sorption process is often described as isotherm
linearity and sorption-desorption singularity. Although these ideal conditions may be true
under certain circumstances, nonideal sorptive behaviors such as isotherm nonlinearity
(Spurlock et al., 1995; Weber and Huang, 1996; Xing and Pignatello, 1997; Xing, 1998)
and sorption-desorption hysteresis (Di Toro and Horzempa, 1982; Miller and Pedit, 1992;
Bowman and Sans, 1985) have been reported. Sorption nonideality can lead to significant
3 variations in the predicted transport of the contaminants in the environment compared to the ideal model (Brusseau and Rao, 1989).
Studies have shown that organic compounds preferentially sorb to soil organic matter (SOM) over the mineral phase of soil in aqueous systems due to strong competition of water molecules for mineral surfaces (Chiou, 1989; Karickhoff, 1984). The dominant components of SOM are humic substances (HS). HS are “a series of relatively high- molecular-weight, yellow to black colored substances formed by secondary synthesis reactions” (Stevenson, 1994). They can be further classified into three subclasses: fulvic acid (FA), humic acid (HA) and humin. FA is the fraction of HS that is soluble in water under all pH conditions; HA is the fraction insoluble in water under acidic conditions (pH <
2) but soluble at higher pH values; humin is the fraction insoluble in water at any pH value
(Aiken et al., 1985). As the principle sorbents in soil, HS play a fundamental role in chemical transport, bioavailability, and toxicity because they govern the distribution of a chemical between bio-accessible, mobile phases (solution) and bio-inaccessible, immobile phases (solids). However, they are one of nature’s least understood materials. Their interactions with organic compounds need to be explored intensively and interpreted on a molecular level.
Generally organic chemicals interact with SOM by two kinds of forces, i.e., specific and nonspecific forces. Specific forces include electrostatic force, covalent bond, charge transfer and hydrogen bond. Nonspecific forces are mainly van der Waals forces. Specific forces are the driving forces for sorption of ionic and polar compounds. Sorption of nonpolar compounds often involves nonspecific forces.
4 Sorption of nonpolar compounds by SOM was often considered as a partitioning process, i.e., the sorbed material dissolves into the solid phase by forces common to solution (Chiou et al., 1979, 1983). Sorption data fits a linear equation:
qe = (1.1)
where ^e(pg/g) is the mass of solute sorbed per unit mass of solid, Cg (pg/mL) is the solute concentration in solution, and Kj is the distribution coefficient. Kj can be related to the well-known octanol-water partitioning coefficient, Kow, and the mass fraction of organic carbon, foe, by an empirical equation:
log(KjJ foe) = log Koc = a log Kow + h (1.2)
The coefficients a and b are obtained experimentally and relate to sorbent properties
(Karickhoff et al., 1984; Hassett et al., 1983).
Evidence supporting the partition mechanism includes: (i) Equilibrium i.sotherms
are linear; (ii) Equilibrium heats of sorption for solutes are less exothermic than heats of
solute condensation from water; (iii) The system shows lack of solute competition (Chiou,
1989).
Although the partition model works well for some systems, mounting facts have
shown deviations from this simple model, including nonlinear isotherms (Spurlock et al.,
1995; Weber and Huang, 1996; Xing and Pignatello, 1997; Xing, 1998), competitive
5 effects in multisolute systems (Pignatello, 1991; Xing et al., 1996; Xing and Pignatello,
1997), and concentration-dependent enthalpy of sorption (Young and Weber, 1995). These
findings are exactly opposite to the evidence used to support the partitioning model and,
moreover, imply a non-uniform sorption potential.
To account for this discrepancy from the simple partitioning model, multi-sites
models have emerged in the literature, notably the distributed reactivity model (LeBoeuf
and Weber, 1997, references therein) and the dual-mode model (Xing and Pignatello,
1996; Xing et al., 1996). In the distributed reactivity model, sorbate molecules access
different regions or domains of a particle via different sorption mechanisms and the
composite isotherm is the result of a series of near-linear partition reactions and nonlinear
adsorption reactions (Weber and Huang, 1996). In the dual-mode model, the solid phase of
SOM is postulated to contain both a partitioning (Henry’s law) domain comprised of
amorphous SOM and an adsorption-like (Langmuir) domain comprised of isolated fixed
“sites” within the matrix. Total sorption is thus the sum of partition and Langmuir terms:
Re = KpC, + X (1.3) /=1 1 +
where ge is the total sorbed concentration (pg/g), Q is the solution phase concentration
(pg/mL), Kp is the partition domain coefficient, and and g ° are the Langmuir affinity and
capacity constant, respectively, for each of n different sites.
6 In practice, these multi-sites models were often expressed by an empirical
Freundlich equation, which can be thought as a result of superposition of partitioning and
Langmuir terms in eq 1.3:
q, = K,C,^ (1.4)
where qe and Ce are defined as before, and Kf and N are constants. N is also known as an index of isotherm nonlinearity (Weber et al., 1992). The closer to unity is the TV value, the more linear is the isotherm.
The two-domain model was first introduced to describe the combined rubbery and glassy state sorption behavior of chemically homogeneous polymers (Vieth and Sladek,
1965). Polymers can be classified as rubbery or glassy on the basis of internal polymer segment motion, void space, and cohesive forces. Rubbery polymers have a relatively expanded, flexible structure; while glassy polymers have a more rigid, condensed structure.
Sorption in rubbery polymers obeys Henry’s law and is noncompetitive; while sorption in glassy polymers is nonlinear and competitive. It was suggested that Langmuir sites postulated to be unrelaxed free volume (“microvoids”) exist in rigid regions of the glassy polymer matrix. SOM sorption behavior has been analogized to that of glassy polymers by recent investigators (Weber and Huang, 1996; Xing et al., 1996). In their hypotheses, SOM contains an amorphous domain and a condensed domain. Using X-ray diffraction (XRD) technique, several researchers have reported the presence of condensed aromatic and expanded aliphatic regions in humic substances from various locations (Schnitzer et al..
7 1991; Xing and Chen, 1999). Brusseau (1993) analyzed the functional dependence of equilibrium and nonequilibrium sorption parameters on solute molecular descriptors
(including molecular surface area, van der Waals volume and molecular connectivity) for
29 organic compounds (e.g., polycyclic aromatic hydrocarbons, chlorobenzenes, etc.) and two soils using the quantitative structure-activity relationship approach and found that the correlation patterns exhibited by the soil data were similar to the patterns exhibited by the polymer systems. Hence, the polymer analogue may be appropriate for both equilibrium and nonequilibrium sorption by soil.
SOM compositions also have great impact on sorption of organic compounds. In the investigation of naphthalene sorption on a set of soil samples varying in age, diagenesis, and composition of organic matter, Xing (1997) found that polarity and aromaticity of
SOM significantly influenced uptake of naphthalene by soils. Organic carbon normalized sorption coefficient {Kqc) varied inversely with effective polarity and directly with aromaticity. Huang and Weber (1997) examined phenanthrene sorption equilibria for 10 natural sorbents ranged in geological ages and organic matter compositions and concluded that the sorption affinities as well as isotherm nonlinearities for phenanthrene correlate inversely with the 0/C atomic ratios of the SOM. These results can be explained successfully in the context of the two-domain model.
Research on interactions of dissolved organic matter (DOM) with nonionic organic compounds has shown that environmental variables such as ionic strength, pH, and type of cations may influence the extent of binding. Carter and Suffet (1982) observed a decrease in the binding of DDT by humic acids when the pH was raised from 6.0 to 9.2. At pH 8.3, either the addition of calcium or an increase in ionic strength increased binding. Traina et
8 al. (1989) studied the influence of pH, ionic strength, and cation types on the binding of naphthalene by water-soluble soil organic matter extracted from a muck soil and observed neither pH (1.5-7.3) nor ionic strength (0.05-0.5 mol L*') effects. However, the presence of
AP‘" appeared to decrease the binding of naphthalene relative to the amount bound in the presence of Na^ and Ca^^. Schlautman and Morgan (1993) investigated the influence of solution chemistry on the binding of anthracene, pyrene and perylene by well-characterized humic material and found that the pH effects agree with the findings of Carter and Suffet
(1982), while the ionic strength effects are opposite. The magnitude of aqueous chemistry effects also depends on the relative sizes of the solute and humic substance. These results
were explained by the configuration changes of DOM under different aqueous chemistry
environments. DOM is a complex, macromolecular polyelectrolyte containing a wide
variety of functional groups such as carboxylic and phenolic groups and is typically anionic
in soil solutions. At high pH and low ionic strength, DOM usually exists as well ionized,
easily deformable linear polymers. In contrast, humic polymers may coil as the pH
decreases or as the ionic strength increases (Ghosh and Schnitzer, 1980). Since
contradictory results have been found from these studies, caution should be taken to
designate a proposed mechanism based on sorption data alone.
Desorption data could provide additional evidence to support or reject a
hypothesized mechanism. The existence of hysteresis in desorption of polyaromatic
hydrocarbons (Kan et al., 1994; Huang and Weber, 1997; Weber and Huang, 1998)
contradicted the prediction of partitioning mechanism, but were consistent with the two-
domain models. Irreversible binding and physical trapping are two major causes for true
hysteresis, in which some fraction of sorbate could not be recovered even with extensive
9 extraction procedures (Brusseau and Rao, 1989; Weber et al., 1998). Irreversible binding includes covalent bond or other specific forces (e.g., electrostatic force) between the organic contaminants and SOM. Large activation energy is needed to overcome the irreversible binding to bring the sorbate into the aqueous phase. Physical trapping refers to the process that certain sorbates diffuse into micropores within the SOM matrix and sterically resist coming out.
From the above literature review, we can see that sorption nonlinearity appears to
be a major concern for the selection of appropriate equilibrium models. Although the
effects of metal cations on the binding of hydrophobic compounds by dissolved organic
matter have been studied, little has been done by particulate humic substances. Therefore,
this study examined whether there are changes in sorption nonlinearity when soil humic
substances were saturated with various metal cations, and if so, how these changes affect
sorption-desorption reversibility.
Objectives
1. To explore the effects of different saturating cations on sorption/desorption of nonionic
organic compounds by humic substances.
2. To assess the effects of varying molecular size and polarity of the sorbing compound on
sorption/desorption.
3. To conduct an analysis of site-energy distribution for sorption data.
10 References
Aiken, G.R., D.M. McKnight, R.L. Wershaw, and P. MacCarthy. 1985. An introduction to humic substances in soil, sediment, and water. In Humic substances in soil, sediment, and water: Geochemistry, isolation and characterization. G.R. Aiken, P. MacCarthy, R.I. Malcolm, and R.S. Swift (eds.). Wiley-Interscience, New York. pp. 1-9.
Bowman, B.T., and W.W. Sans. 1985. Partitioning behavior of insecticides in soil-water systems: II. Desorption hysteresis effects. J. Environ. Qual. 14:270-273.
Brusseau, M.L. 1993. Using QSAR to evaluate phenomenological models for sorption of organic compounds by soil. Environ. Toxicol. Chem. 12:1835-1846.
Brusseau, M.L., and P.S.C. Rao. 1989. Sorption nonideality during organic contaminant transport in porous media. Crit. Rev. Environ. Control. 19:33-99.
Carter, C.W., and I.H. Suffet. 1982. Binding of DDT to dissolved humic materials. Environ. Sci. Technol. 16:735-740.
Carter, M.C., J.E. Kilduff, and W.J. Weber, Jr. 1995. Site energy distribution analysis of preloaded adsorbents. Environ. Sci. Technol. 29:1773-1780.
Chiou, C.T. 1989. Theoretical considerations of the partition uptake of nonionic organic compounds by soil organic matter. In Reactions and movement of organic chemicals in soil. B.L. Sawhney, and K. Brown (eds.). Soil Science Society of America, Madison, WI. pp. 1-29.
Chiou, C.T., L.J. Peters, and V.H. Freed. 1979. A physical concept of soil-water equilibria for nonionic organic compounds. Science (Washington, DC). 206: 831-832.
Chiou, C.T., P.E. Porter, and D.W. Schmedding. 1983. Partition equilibria of nonionic organic compounds between soil organic matter and water. Environ. Sci. Technol. 17:227-231.
Di Toro, D.M., and L.M. Horzempa. 1982. Reversible and resistant component of PCB adsorption-desorption isotherms. Environ. Sci. Technol. 16:594-602.
Ghosh, K., and M. Schnitzer. 1980. Macromolecular structures of humic substances. Soil Sci. 129:266-276.
11 Hassett, J.J., W.L. Banwart, and R.A. Griffin. 1983. Correlation of compounds properties with sorption characteristics of nonpolar compounds by soils and sediments: concepts and limitations. In Environment and solid wastes. C.W. Francis, and S.l. Auerbach (eds.). Butterworths, Boston, pp. 161-178.
Huang, W., and W.J. Weber, Jr. 1997. A distributed reactivity model for sorption by soils and sediments. 10. Relationships between desorption, hysteresis, and the chemical characteristics of organic domains. Environ. Sci. Technol. 31:2562-2569.
Kan, A.T., G. Fu, and M.B. Tomson. 1994. Adsorption/desorption hysteresis in organic pollutant and soil/sediment interaction. Environ. Sci. Technol. 28: 859-867.
Karickhoff, S.W. 1984. Organic pollutant sorption in aquatic systems. J. Hydraulic. Engineer. 110:707-735.
Miller, C.T., and J.A. Pedit. 1992. Use of reactive surface-diffusion model to describe apparent sorption-desorption hysteresis and abiotic degradation of lindane in a subsurface material. Environ. Sci. Technol. 26:1417-1427.
Pavlostathis, S.G., and G.N. Mathavan. 1992. Desorption kinetics of selected volatile organic compounds from field contaminated soils. Environ. Sci. Technol. 26:532- 538.
Schlautman, M.A., and J.J. Morgan. 1993. Effects of aqueous chemistry on the binding of polycyclic aromatic hydrocarbons by dissolved humic materials. Environ. Sci. Technol. 27:961-969.
Schnitzer, M., H. Kodama, and J.A. Ripmeester. 1991. Determination of the aromaticity of humic substances by X-ray diffraction analysis. Soil Sci. Soc. Am. J. 55:745-750.
Spurlock, F.C., K. Huang, and M.Th. van. Genuchten. 1995. Isotherm nonlinearity and nonequilibrium sorption effects on transport of fenuron and monuron in soil columns. Environ. Sci. Technol. 29:1000-1007.
Steinberg, S.M., J.J. Pignatello, and B.L. Sawhney. 1987. Persistence of 1,2-dibromoethane in soils: Entrapment in intraparticle micropores. Environ. Sci. Technol. 21:1201- 1208.
Stevenson, F.J. 1994. Chapter 2, Extraction, fractionation, and general chemical composition of soil organic matter. In Humus Chemistry. John Wiley & Sons, Inc. New York. pp. 33.
Traina, S.J., D.A. Spontak, and T.J. Logan. 1989. Effects of cations on complexation of naphthalene by water-soluble organic carbon. J. Enviroa Qual. 18:221-227.
Vieth, W.R., and K. J. Sladek. 1965. A model for diffusion in a glassy polymer. J. Colloid Sci. 20:1014-1033.
12 Weber, W.J. Jr., W. Huang, and H. Yu. 1998. Hysteresis in the sorption and desorption of hydrophobic organic contaminants by soils and sediments. 2. Effects of soil organic matter heterogeneity. J. Contam. Hydrol. 31:149-165.
Weber, W.J., Jr., and W. Huang. 1996. A distributed reactivity model for sorption by soils and sediments. 4. Intraparticle heterogeneity and phase-distribution relationships under nonequilibrium conditions. Environ. Sci. Technol. 30: 881-888.
Weber, W.J., Jr., P.M. McGinley, and L.E. Katz. 1992. A distributed reactivity model for sorption by soils and sediments. 1. Conceptual basis and equilibrium assessments. Environ. Sci. Technol. 26:1955-1962.
Xing, B. 1997. The effect of the quality of soil organic matter on sorption of naphthalene. Chemosphere. 35:633-642.
Xing, B. 1998. Reaction of toluene with soil organic matter. J. Environ. Sci. Health. B33:293-305.
Xing, B., and J.J. Pignatello. 1996. Time-dependent isotherm shape of organic compounds in soil organic matter: Implications for sorption mechanism. Environ. Toxicol. Chem. 15, 1282-1288.
Xing, B., and J.J. Pignatello. 1997. Dual-mode sorption of low-polarity compounds in glassy poly (vinyl chloride) and soil organic matter. Environ. Sci. Technol. 31:792- 799.
Xing, B., and Z. Chen. 1999. Spectroscopic evidence for condensed domains in soil organic matter. Soil Sci. 164:40-47.
Xing, B., J.J. Pignatello, and B. Gigliotti. 1996. Competitive sorption between atrazine and other organic compounds in soils and model sorbents. Environ. Sci. Technol. 30:2432-2440.
Young, T.M., and W.J. Weber, Jr. 1995. A distributed reactivity model for sorption by soils and sediments. 3. Effects of diagenetic processes on sorption energetics. Environ. Sci. Technol. 29:92-97.
13 CHAPTER 2
EFFECTS OF METAL CATIONS ON SORPTION/DESORPTION OF ORGANIC
COMPOUNDS BY HUMIC ACIDS
Abstract
Sorption/desorption of selected model compounds including naphthalene, phenanthrene and a-naphthol by cation-saturated humic acids (HA) was investigated using batch equilibration method. All sorption isotherms were found to be nonlinear with
Freundlich exponents (A values) less than one. values for all tested compounds were almost unchanged in Ca-HA, but significantly decreased in Al-HA as compared to untreated HA. For the same sorbent, A values followed the order: naphthalene > phenanthrene > a-naphthol. Hysteresis occurred in all systems excluding naphthalene-HA and naphthalene-Ca-HA. The magnitude of hysteresis, evaluated by the ratio of the N values for desorption and sorption, followed the order: HA ~ Ca-HA < Al-HA and naphthalene < phenanthrene < a-naphthol. There was a good correlation (r > 0.90) between the magnitude of hysteresis and the values of sorption isotherms. The results from this study suggested that properties of both the sorbates and the sorbents may influence sorption mechanisms. The results could not be explained by the well-known partitioning mechanism alone, but were compatible with the dual-mode model, in which hole-filling concurs with solid-phase partitioning. This research was expected to shed light on the future practice of cleaning up contaminated sites.
14 Introduction
Humic substances (HS) are ubiquitous in waters, soils and sediments. They are “a series of relatively high-molecular-weight, yellow to black colored substances formed by secondary synthesis reactions” (Stevenson, 1994). HS play an important role in retaining soil moisture, buffering soil acidity and binding metals, molecules and ions. Therefore, they have great impact on agricultural and environmental activities. Unfortunately, HS are one of nature’s least understood materials considering their heterogeneity and purification difficulty.
Sorption of organic contaminants (pesticides, petrochemicals, anthropogenic wastes, etc.) by HS is an underlying process in determining the transport and fate of these contaminants in the environment. Reliable estimations of sorption parameters are crucial
for the transport modeling. Sorption mechanisms were often derived from sorption data
alone (Chiou, 1989). However, one set of data could be fitted with several models (Carter
et al., 1995). Therefore, other evidence needs to be explored to strengthen the hypothesized
mechanisms. To this end, sorption as well as desorption experiments were performed with
model compounds (naphthalene, phenanthrene, and a-naphthol) by HS. The effects of
cations on sorption/desorption mechanisms were also investigated. The results could have
some benefits for remediation practice.
15 Materials and Methods
Sorbates
[Ring-UL-‘^C] naphthalene, phenanthrene and a-naphthol were obtained from
Sigma Chemical Company, and unlabeled compounds were from Aldrich Chemical
Company. The properties of these model compounds were listed in Table 2.1.
Humic acid extraction and purification
The humic acid extraction and purification followed the procedures proposed by
Schnitzer (1982). Soil samples were a peat soil (Oxyaquic Hapludalfs, SOM 57% by weight) from Chelsea, MI and a mineral soil (Typic Fragiochrept, SOM ~10%) from
Amherst, MA. Soil was air-dried and passed through a 2 mm sieve. Then soil and 0.1 M
Na4P207 solution were put together into a 1 L polypropylene flask (solid to solution ratios were 3 g/100 mL for peat and 10 g/100 mL for soil). The mixture was shaken for 24 h at room temperature after displacing air from the flask with Nj gsis. The soil suspension was centrifuged at 1,000 g for 10 min and the supernatant was stored in a large beaker, 'fhe precipitate was extracted with Na4P207 solution one more time. After centrifugation, the supernatant was added to the previously stored one. The supernatant was then acidified to pH 1.5 with concentrated HCl and allowed to settle for 24 h. Coagulated HA was then separated by centrifuging at 7,000 g for 10 min. One gram of HA was mixed with
16 Table 2.1 Properties of model compounds 17 100 mL HCl/HF solution (0.5 mL of concentrated HCl and 0.5 mL of 49% HF in 99 mL of deionized water) and shaken for 24 h at room temperature. HA was removed from the mixture by centrifugation, washed with deionized water and freeze-dried for subsequent use.
Elemental Composition of HA
The composition of HA was determined in the Microanalysis Laboratory of
University of Massachusetts at Amherst, MA. Chelsea HA consists of 51.78% C, 4.17% H,
40.19% O, 2.96% N, and 0.9% ash. Amherst HA consists of 42.09% C, 3.84% H, 42.58%
O, 3.69%N, and 7.8% ash.
Metal cation saturation of HA
Thirty mL each of 0.1 M CaClj and 0.1 M AICI3 solutions was separately added into one of two centrifuge tubes containing 5 g HA sample. The tubes were placed onto a mixer for 48 h and then centrifuged at 1,000 g for 15 min. After decanting the supernatant, the
HA sample was re-mixed with fresh salt solution and centrifuged. The above steps were repeated three more times. The equilibration period for the last two runs was 15 min. The sample was then washed with deionized water and precipitated by centrifuging at 10,000 g for three times to remove the excess salt. The cation-saturated HA was then freeze-dried.
18 Sorption experiments
Sorption experiments were conducted using batch equilibration method (Xing,
1996a) in 8-mL screw-cap vials (minimal headspace) with Teflon-lined septa for naphthalene and a-naphthol, and 40-mL vials for phenanthrene. The sorbents were untreated HA, Ca- and Al-HA. The background solution was 0.01 M CaClj (0.01 M AICI3
for Al-HA) and 200 pg/mL HgCl2 as a biocide. The ’"C-labeled chemicals along with their
nonradioactive stock solutions were mixed with humic acids at different solid to solution
ratios depending on their water solubilities and affinities with HA. The initial
concentrations were evenly spaced in log-scale and ranged from 0.008 to 15 pg/mL for
naphthalene, 0.006 to 0.8 pg/mL for phenanthrene, and 0.1 to 316 pg/mL for a-naphthol.
Duplicates were run for naphthalene and a-naphthol while only one replicate was run for
phenanthrene. Blanks without HA were run for each initial concentration. The HA
suspensions were shaken on hematology mixers giving rocking-rotating motions at 23 ± 1
°C for 3 d. A three-day period was considered to be long enough to achieve apparent
sorption equilibrium because preliminary sorption experiments for up to 7 days have shown
no significantly increasing sorption after 1 d. The vials were centrifuged at 1,000 g for 20
min, and about 1 mL of the supernatant was sampled for liquid scintillation counting
(Beckman LS 3801). Because of little sorption by the vials, the amount of test compound
sorbed by HA was calculated by the mass difference.
19 Desorption experiments
Desorption was conducted by sequential decant-refill steps following the completion of sorption. At the end of sorption experiments, solids were separated from the aqueous solutions by centrifugation at 1,000 g for 20 min and an aliquot (~ 1 mL) of supernatant was withdrawn from each vial for liquid-scintillation counting. Then about 6 mL (30 mL for phenanthrene) of the remaining supernatant was replaced with the same volume plus 1 mL of the background solution (i.e., 7 mL for naphthalene and a-naphthol, and 31 mL for phenanthrene). The dilution volumes were determined by weight. After dilution, the vials were shaken on the hematology mixers for 3 days. The suspensions were then centrifuged and an aliquot of supernatant was extracted for analysis. The remaining supernatant was again replaced by the same volume of the background solution and the above process repeated three more cycles. The concentration of the test '^’C sorbate present in the supernatant solution after each desorption was determined by liquid scintillation counting. Desorbed '"‘C sorbate was then calculated at each desorption stage. The amount of sorbate remaining on HA at each desorption stage was calculated as the difference between the initial sorbed amount and the desorbed amount.
Data analysis
All sorption/desorption data were fitted to the logarithmic form of Freundlich equation:
log = log a:/-+ A log Q (2.1)
20 where qe and Cg are sorbed (|ig/g) and solution (|ig/mL) concentration at sorption/desorption, respectively. Kp and N are Freundlich constants for sorption/desorption. Equilibrium isotherms were plotted by log qe vs. log Q. log K[. and N were obtained from the intercept and slope of the plot, respectively.
The ratio of Freundlich exponents for desorption and sorption, i.e., (subscripts d and s refer to desorption and sorption, respectively), has been considered as the hysteresis index by several authors (O’Connor et al., 1980; Barriuso et al., 1994). If sorption/desorption of a chemical is nonhysteretic, NdlNs=\. The N^Ns values were calculated for all the systems in this study.
Results and Discussions
Sorption
Sorption data fitted well with the Freundlich model (r^ > 0.999, Table 2.2). The calculated Kqc values from isotherm parameters at different equilibrium concentrations
for all test chemicals were listed in Table 2.2. It was obvious that A^oc values are
concentration dependent. All test chemicals exhibited greater affinity for SOM at lower
equilibrium concentrations. This was also indicated by N values for sorption, which are
less than one. The presence of a small amount of high-energy sites in the SOM may be
responsible for this greater affinity at lower solution concentrations (Xing, 1998). As the
solution concentration increased, these high-energy sites may be easily saturated,
therefore the distribution coefficients dropped.
21 Table 2.2 Sorption parameters and calculated Koc values for Amherst HA B 00 •4-> CAl (D B O O' O' 'ci) '5i) 1—I 3 II II zL s 'm 'ob 3 3 g • ^ "cS C (U G IT) o o O CM o o o m o o o m r—1 (N m o o o m o X rn CN < -H -H < phenanthrene < naphthalene, indicating that the degree of sorption nonlinearity was in the order: a-naphthol > phenanthrene > naphthalene. This result could be expected by the dual-mode sorption model. Since a-naphthol is a polar compound, while phenanthrene and naphthalene are nonpolar, there would be more specific sites for a-naphthol sorption. Therefore sorption of a-naphthol exhibited the greatest degree of nonlinearity. Molecular size effect would be responsible for the difference between sorption of naphthalene and phenanthrene. Phenanthrene molecule is bigger, therefore more easily entrapped in micropores within SOM than naphthalene molecule. Table 2.2 also showed that the introduction of A1 to HA significantly reduce sorption values, while Ca has little effect on values. It has been plausibly conjectured that sorption of hydrophobic organic compounds by physically condensed SOM exhibits a greater degree of nonlinearity than sorption by amorphous SOM (Huang and Weber, 1997; Xing and Pignatello, 1997). Saturating HA with cations would cause the macromolecules to coagulate and become more condensed due to the cation bridges between the anionic functional groups of HA (Ghosh and Schnitzer, 1980; Traina et al., 1989; Schlautman and Morgan, 1993). In this configuration changing process, trivalent ions (e.g. Al^^) are more effective coagulating agents than divalent ions (e.g. Ca^^). Data from Table 2.2 were in accordance with the above phenomena. 23 Desorption Desorption isotherms for all systems were constructed using consecutive desorption data corresponding to three different initial concentrations and were illustrated in Fig. 2.1 through Fig. 2.9. Each point on the isotherm represented the average of the duplicate. The data also fitted well with the Freundlich model (r^ ranged from 0.909 to 1, Table 2.3). Hysteresis indices were calculated by the ratios of Freundlich exponents for desorption and sorption at three initial concentrations and listed in Table 2.3. An alternative method of quantifying hysteresis with the relative difference of solid-phase solute concentrations at fixed residual solution concentration has been proposed by Huang et al. (1998). In their method, they only used the first-cycle desorption isotherm constructed by connecting the first step desorption points from each initial concentration. However, strictly speaking, it is improper to use the first-cycle desorption isotherm to quantify hysteresis because, in the presence of nonreversibility, the desorption points obtained depend on the previous starting points and the details of the desorption procedure, such as the dilution factor (Bowman and Sans, 1985). Chemicals were supposed to be sorbed by SOM with varying sorption energies (Xing et al., 1996a, 1996b). For the first desorption step, it is likely that sorbed molecules with the lowest sorption energies first come into solution. A significant decrease in percent desorbed during subsequent desorption steps was reported (Grover, 1975). This was mainly due to the increasing desorption resistance of the remaining solid-phase solute. If only the first desorption data were considered, some systems would not have exhibited hysteresis, but the same systems showed significant hysteresis after several desorption 24 Table 2.3. Desorption parameters for chemicals on Amherst HA X < < CJ X < X < Initial 1 L CJD u JZ CO 00 s CJ o o c o (U c VO On O ON NO NO o On (N ON ON ON ON o ON ON cn O O On r- NO 00 d> o (N fN m O p O m o ON O'l m CN o (N oo o ON r- o p o (N cd CON CON (d 1—1 oo cd m 9V0 1—( ON in (d NO o cd CN (d p 1—1 oo (ON (d cd NO cn cn m m 1—c r- 25 , • oo in cd ON o (d 00 100 • ‘B 'So hJ 'Si) 'So w X3 c/3 G =L e G c C/2 :± =L ra Sorbed Concentration Sorbed Concentration Figure 2.1SuccessivedesorptionofnaphthalenebyAmherstHA Figure 2.2Successivedesorptionof naphthalene byAmherstCa-HA 26 Sorbed Concentration Sorbed Concentration Figure 2.3SuccessivedesorptionofnaphthalenebyAmherstAl-HA Figure 2.4Successivedesorptionof phenanthrene byAmherstHA 27 Sorbed Concentration Sorbed Concentration Figure 2.5 Figure 2.6Successivedesorptionof phenanthrene byAmherstAl-HA Successive desorptionofphenanthrenebyAmherstCa-HA 28 Sorbed Concentration Sorbed Concentration 1 e+5 1 e+1-V Figure 2.8Successivedesorptionof a-naphthol byAmherstCa-HA Figure 2.7Successivedesorptionofa-naphtholbyAmherstHA 0.01 0.01 0.1110 100 1000 sorption 0.1 110100 Solution Concentration(ng/mL) Solution Concentration(^ig/rnL) 29 1000 1 e + 5 o sorption desorption, 0.32 [xglml cz o ra a> o c: o o> O oi ■o a a> o c/5 1 e +1 0.01 0.1 1 10 100 1000 Solution Concentration (lag/mL) Figure 2.9 Successive desorption of a-naphthol by Amherst Al-HA steps (Fig. 2.4). Similar findings have been reported by Bowman and Sans (1977). Therefore Huang’s method was not used in this work. It was visually observed from Fig. 2.1 and Fig. 2.2 that naphthalene showed no hysteresis at all in HA and Ca-HA, however, it did show hysteresis in Al-HA (Fig. 2.3). Examining the data in Table 2.4, we found that the hysteresis for phenanthrene in HA and Ca-HA were comparable, while more hysteresis was found in Al-HA. The hysteresis for a- naphthol were much greater than that for naphthalene and phenanthrene. From Table 2.4, we could also see a trend of hysteresis decreasing with increasing the initial solute concentrations. This may be due to less resistant fractions present at high concentration than at low concentration. 30 Table 2.4 Hysteresis indices for chemicals in Amherst HA initial concn Chemical (pg/mL) HA Ca-HA Al-HA Naphthalene 0.024 1.03 1.06 0.75 l.O 1.00 1.00 0.75 8.0 1.03 0.87 0.73 Phenanthrene 0.030 0.68 0.62 0.36 0.090 0.72 0.65 0.43 0.46 0.71 0.67 0.50 a-Naphthol 0.32 0.25 0.25 0.22 10 0.36 0.41 0.27 100 0.49 0.55 0.31 Potential causes for sorption/desorption hysteresis have been thought to include (Weber et ah, 1998): (i) experimental artifacts, such as solute loss due to volatilization and/or sorption to the vials, (ii) physical entrapment of the solute, or (iii) site-specific binding between the solute and the sorbent. The first factor can be excluded from this study because actually no hysteresis was observed for naphthalene in HA and Ca-HA. As stated earlier, micropores within rigid and condensed SOM domains may behave like Langmuir sites which were capable of trapping the hydrophobic organic solutes without involving specific interactions. Small molecules such as naphthalene may move freely through micropores thus exhibiting little or slight hysteresis, whereas large molecules such as phenanthrene could be trapped into micropores and were resistant to desorption. 31 Meanwhile, more hysteresis occurred in the relatively more condensed (i.e., cross-linked) Al-HA than in HA or Ca-HA. In addition to being physically entrapped within micropores in SOM, a-naphthol could undergo oxidative coupling with SOM (Burgos et al., 1996). The contribution of this irreversible binding could help to explain the more resistant fractions in a-naphthol than in naphthalene and phenanthrene. It is interesting to note that the hysteresis indices increased significantly with the initial concentration for a-naphthol in HA and Ca-HA, while slight increase was observed in Al-HA. This was because that hysteresis originated from two sources: site-specific bonding and physical trapping. Since the molecular size of a-naphthol is about the same as that of naphthalene, we could assume that physical trapping of a-naphthol will be similar to that of naphthalene. Therefore, in desorption of a-naphthol in HA and Ca-HA, only the covalent bonding contributed to the hysteresis. At higher concentration, the irreversible binding sites may be easily saturated and no additional source will contribute to hysteresis, resulting a remarkable decrease of resistant fractions to desorption. In the case of Al-HA, physical trapping will still contribute to hysteresis after the irreversible binding sites is saturated, therefore the hysteresis indices only increased slightly with the initial concentration. Figure 2.10 through Figure 2.12 showed a good correlation (r'> 0.90) between sorption nonlinearity and the hysteresis index for desorption of the three test compounds in Amherst HA. This correlation was anticipated because both sorption nonlinearity and hysteresis have the same origins, i.e., irreversible bonding and physical trapping. Since sorption/desorption process for many organic compounds is often very slow (in weeks or 32 months), it is impractical to conduct consecutive desorption experiments for such compounds. The above correlation would facilitate us to reasonably estimate the magnitude of hysteresis in a given soil system based on sorption parameters. Figure 2.10 Correlation of sorption nonlinearity (Ns) with hysteresis index (NJNs) at the highest initial concentrations for the three chemicals 1.1 1 1 .0 - • naphthalene • • ■ phenanthrene X 0.9 - ^ a-naphthol O) X) 0.8 ■ ■ c/> 0.7 - in 0.6 - 0.5 - 0.4 - ▲ ▲ r*=0.93 0.3 - ▲ 0.2 1 — 1 1 ^-----1-r - • 0.80 0.82 0.84 0.86 0.88 0.90 0.92 0.94 0.96 Sorption Nonlinearity {N Figure 2.11 Correlation of sorption nonlinearity (Ns) with hysteresis index (NJNs) at the intermediate initial concentrations for the three chemicals 33 1 .2 • naphthalene ■ phenanthrene • • 1.0 ^ a-naphthol 0.8 • ■ 0.6 0.4 ■ ▲ A ▲ 0.2 r*=0.90 0.0 0.80 0.82 0.84 0.86 0.88 0.90 0.92 0.94 0.96 t Sorption Nonlinearity (A/j) Figure 2.12 Correlation of sorption nonlinearity (Ns) with hysteresis index (NJNs) at the lowest initial concentrations for the three chemicals Sorption/desorption for naphthalene in Chelsea HA showed similar results as on Amherst HA. No hysteresis was observed in HA and Ca-HA, but hysteresis did occur in Al-HA at lower solution concentration (Table 2.5). Table 2.5 Hysteresis Indices for naphthalene in Chelsea HA Initial concn (pg/mL) HA Ca-HA Al-HA 0.024 1.03 0.98 0.84 1.0 1.02 0.96 0.85 8.0 0.93 0.95 0.99 34 In conclusion, nonlinear sorption and hysteresis could be successfully interpreted in the context of the dual-mode model. Micropores and specific sorption sites within SOM could well be responsible for both nonlinearity and hysteresis. The results from this study are potentially quite useful for predicting the transport and fate of hydrophobic organic compounds in the environment. References Barriuso, E., D.A. Laird, W.C. Koskinen, and R.H. Dowdy. 1994. Atrazine desorption from smectites. Soil. Sci. Soc. Am. J. 58:1632-1638. Bondi, A. 1964. Van der Waals volumes and radii. J. Phys. Chem. 68:441-451. Bowman, B.T., and W.W. Sans. 1977. Adsorption of parathion, fenitrothion, methyl parathion, aminoparathion and paraoxon by Na^, Ca^^, and Fe^^ montmorillonite suspensions. Soil Sci. Soc. Am. J. 41:514-519. Bowman, B.T., and W.W. Sans. 1985. Partitioning behavior of insecticides in soil-water systems: II. Desorption hysteresis effects. J. Environ. Qual. 14:270-273. Burgos, W.D., J.T. Novak, and D.F. Berry. 1996. Reversible sorption and irreversible binding of naphthalene and a-naphthol to soil: Elucidation of processes. Environ. Sci.Technol. 30:1205-1211. Chiou, C.T. 1989. Theoretical considerations of the partition uptake of nonionic organic compounds by soil organic matter. In Reactions and movement of organic chemicals in soil. B.L. Sawhney, and K. Brown (eds.). Soil Science Society of America, Madison, WI. pp. 1-29. Ghosh, K., and M. Schnitzer. 1980. Macromolecular structures of humic substances. Soil Sci. 129:266-276. Grover, R. 1975. Adsorption and desorption of urea herbicides on soils. Can. J. Soil Sci. 55:127-135. Howard, P.H., and W.M. Meylan (eds.). 1997. Handbook of Physical Properties of Organic Chemicals. Lewis publishers. 35 Huang, W., and WJ. Weber, Jr. 1997. A distributed reactivity model for sorption by soils and sediments. 10. Relationships between desorption, hysteresis, and the chemical characteristics of organic domains. Environ. Sci. Technol. 31:2562-2569. Huang, W., H. Yu, and W.J. Weber, Jr. 1998. Hysteresis in the sorption and desorption of hydrophobic organic contaminants by soils and sediments. 1. A comparative analysis of experimental protocols. J. Contam. Hydrol. 31:129-148. O’Connor, G.A., P.J. Wierenga, H.H.Cheng, and K.G. Doxtader. 1980. Movement of 2,4,5- T through large soil columns. Soil Sci. 130:157-162. Pavlostathis, S.G., and G.N. Mathavan. 1992. Desorption kinetics of selected volatile organic compounds from field contaminated soils. Environ. Sci. Technol. 26:532- 538. Schlautman, M.A., and J.J. Morgan. 1993. Effects of aqueous chemistry on the binding of polycyclic aromatic hydrocarbons by dissolved humic materials. Environ. Sci. Technol. 27:961-969. Schnitzer, M. 1982. Chapter 30, Organic matter characterization. In Method of soil analysis (Part 2). A.L. Page, R.H. Miller, and D.R. Keeney (eds.). American Society of Agronomy, Soil Science Society of America, Madison, WI. pp. 581-585. Stevenson, F.J. 1994. Chapter 2, Extraction, fractionation, and general chemical composition of soil organic matter. In Humus Chemistry. John Wiley & Sons, Inc. New York. pp. 33. Traina, S.J., D.A. Spontak, and T.J. Logan. 1989. Effects of cations on complexation of naphthalene by water-soluble organic carbon. J. Enviroa Qual. 18:221-227. Weber, W.J. Jr., W. Huang, and H. Yu. 1998. Hysteresis in the sorption and desorption of hydrophobic organic contaminants by soils and sediments. 2. Effects of soil organic matter heterogeneity. J. Contam. Hydrol. 31:149-165. Xing, B., and J.J. Pignatello. 1996. Time-dependent isotherm shape of organic compounds in soil organic matter: Implications for sorption mechanism. Environ. Toxicol. Chem. 15, 1282-1288. Xing, B., and J.J. Pignatello. 1997. Dual-mode sorption of low-polarity compounds in glassy poly (vinyl chloride) and soil organic matter. Environ. Sci. Technol. 31:792- 799. Xing, B., J.J. Pignatello, and B. Gigliotti. 1996. Competitive sorption between atrazine and other organic compounds in soils and model sorbents. Enviroa Sci. Technol. 30:2432-2440. 36 CHAPTER 3 SITE-ENERGY DISTRIBUTION ANALYSIS OF ORGANIC CHEMICAL SORPTION BY SOIL ORGANIC MATTER Abstract Sorption of several hydrophobic organic compounds (HOCs) by selected soils as well as their humic substance fractions was examined using batch equilibration method. Part of the sorption data was from a previous paper (Xing et al., 1996). The results could not be explained by the well-known partitioning mechanism alone, but were consistent with the dual-mode sorption model for soil organic matter (SOM), in which both solid- phase dissolution and hole-filling mechanisms take place. The heterogeneous nature of the natural sorbents was demonstrated by site-energy distributions derived from the common Freundlich model. The site-energy distribution analysis is useful for examining and understanding the energetic characteristics of a sorbent. This analysis lends further support for the dual-mode model of sorption to SOM. Introduction Sorption of hydrophobic organic compounds (HOCs) by soils and sediments plays an important role in determining their transport and fate in the environment. Unfortunately, complete understanding of sorption mechanisms has not been reached so 37 far. Soil organic matter (SOM) is the primary sorbent for HOCs in soil (Chiou. 1989). Sorption to SOM is widely believed to occur by partitioning, which is the concentration- independent dissolution of a solute into the three-dimensional network of SOM. The partition mechanism is supported by the evidence of linear isotherms, less exothermic sorption heats than heats of solute condensation from water and lack of solute competition in many systems (Chiou, 1989). However, nonlinear isotherms (Huang and Weber, 1998, references therein; Xing et al., 1996; Xing and Pignatello, 1997) and competitive sorption (McGinley et al., 1993; Xing et al., 1996; Xing and Pignatello, 1998) have been reported. Moreover, calculations of sorption heats from isotherms measured at different temperatures revealed that the sorption process became less exothermic with increased solid-phase loading (Young and Weber, 1995; Spurlock, 1995). This implied a non-uniform sorption potential within SOM. Nonlinear isotherms and competitive sorption were explained using a dual-mode model (Xing et al, 1996; Xing and Pignatello, 1997; Xing, 1998), in which a partition mechanism concurs with a hole-filling (adsorption) mechanism. The total sorption may be written as: ^ q.b.C q =K C +y / ' ^ (3.1) ^ p ^ ^\ + bC i=l i e where qe and Ce are the sorbed and solution equilibrium concentration, respectively; Kp is the partition domain coefficient; 6, and qi° are the Langmuir affinity and capacity constants, respectively, for each unique type of site /. According to the dual-mode model. 38 SOM is proposed to have expanded domains and condensed domains and nonlinear sorption takes place only in the condensed domains. Weber and his colleagues proposed a similar dual SOM domain concept (Weber and Huang, 1996; Huang and Weber, 1998), in which the condensed regions embedded in amorphous SOM are responsible for the nonlinear isotherms and competitive sorption. These models indicate the heterogeneous nature of soil organic matter. Isotherm nonlinearity may be attributed to a distribution of site energies. On a theoretical basis, a set of specific isotherm parameters correspond to a specific site-energy distribution. Changes in sorbent characteristics, i.e., changes in site-energy distributions of sorbent, will result in different sets of isotherm parameters. Conversely, we can obtain valuable information of site-energy distributions from the measurements of sorption parameters in various systems. This information may provide an extra tool for understanding the sorption phenomena. Carter et al. (1995) developed a methodology to determine site energy distributions from isotherm parameters. These parameters were obtained from the experiments designed to determine the preloading effects of 1,2,4-trichlorobenzene on sorption of trichloroethylene by activated carbon. They derived approximate site-energy distribution functions from several isotherm models which are special cases of the so- called generalized Langmuir model: {nim) {hex (3.2) \-^{hCX 39 where is the maximum adsorption capacity, m and n are heterogeneity parameters and is a sorption-energy related constant. They found that site-energy distributions corresponding to both Langmuir-Freundlich model (when n=m in eq 3.2) and common Freundlich model (low-concentration limit of eq 3.2) could explain the decrease in the heterogeneity of surface sites after preloading (i.e., values significantly increase with preloading). This is reflected by a decrease in the number of higher energy sites and an increase in lower energy sites. In this paper, the site-energy distribution analysis was extended to sorption of organic chemicals by natural sorbents to evaluate the energy distributions of sorption sites. Sorption of binary solutes by a mineral soil and sorption of single solutes to different components of SOM as well as to cation-saturated SOM were examined. Theoretical According to the theory of sorption on heterogeneous surfaces, the general isotherm equation may be written as: q^(C^)= ^q^{E,C^)F\E)dE (3.3) where qeiCe) is the overall sorption isotherm holding over all the sorption sites, qh{E, Ce) is the local isotherm (e.g., Langmuir isotherm) holding on each homogeneous patch of sites with a sorption energy E, F(E) is the energy distribution function. The integral is set from 40 zero to infinity for convenience because the actual minimum and maximum sorption energies are not known a priori (Carter et al., 1995). Assuming the local isotherm as Langmuir-type, exact site-energy distributions corresponding to various isotherm models (e.g., GLM and its simplifications) have been derived by solving the integral equation for F{E) through application of a Stieltjes transform (Derylo-Marczewska and Jaroniec, 1984). In order to simplify the solution of the above integral equation, the true local isotherm was replaced by a step-function. This technique was referred to as the condensation approximation (Harris, 1968; Cerofolini, 1974). Carter et al. (1995) used the condensation approximation for determining the site- energy distributions in their work. Under the condensation approximation, the local homogeneous isotherm is expressed as: 0 for E where q\E) is the maximum sorption capacity for a homogeneous patch with sorption energy E. It is assumed that at equilibrium solution concentration C^, sorption sites are saturated if their energy is greater than or equal to F(Ce) and are completely free otherwise (Seidel and Carl, 1989). Therefore F can be regarded as the lowest value of binding energy at which condensation (i.e., sorption) occurs at Cg. The relationship between £* and Cg is 41 determined by minimizing the difference between the given step function and the Langmuir-type isotherm, which gives: C =C5exp(-— (3.5) where Cs is the solute solubility in water, R is the universal gas constant, T is absolute temperature. From eq 3.5, we can easily obtain £* = 0 when Cg = Cs. Here the lowest physically realizable sorption energy, i.e., sorption energy corresponding to Cg = G, is set to zero, comparable to Eg in the paper by Carter et al. (1995). Substituting eq 3.4 into eq 3.3, using the relationship between and Q in eq 3.5, and defining F{E) = q°{E)F{E), we have: (3.6) The approximate site-energy distribution F{E^) can be obtained by differentiating eq 3.6 with respect to FT: (3.7) 42 The above energy distribution F{E^) is not normalized and the area under the distribution curve is equal to the overall maximum sorption capacity of the heterogeneous sorbent, namely Qg\ Yf(E’)dE'= Q’^ (3.8) Choosing the “right” form (e.g., common Freundlich model) of qe{F^) in eq 3.7 is important because almost all the models are limited to a certain concentration range. Although the common Freundlich model equation does not approach a limiting value at high concentration, it enjoys a wide use for sorption data over the low concentration ranges which are of environmental interest. Therefore we employ the common Freundlich model in this paper to perform the site-energy distribution analysis. Common Freundlich model equation can be written as: q =Kf(CY (3.9a) or log q^ = log Kf-\- N log (3.9b) where qe and Ce are sorbed and solution concentration, respectively. Kf is sorption coefficient, N is heterogeneity index (Weber et al., 1992). The closer the N value is to unity, the more linear is the isotherm. 43 Substituting eq 3.5 into eq 3.9a, we get: (3.10) Differentiating eq 3.10 with respect to and combining with eq 3.7 yields: -dq^(E') K^N(C/ F{E ) = ^exp(—) (3.11) dE Eq 3.11 was used for the site-energy distribution analysis for sorption data obtained from this study. This equation is actually the same expression as eq 10 in the paper by Carter et al. (1995). Experimental Sorbents Composition of the sorbents used is presented in Table 3.1. Florida Pahokee peat soil (PPS, a Lithic Medisaprist) was obtained from the International Humic Substance Society (University of Minnesota). Pahokee peat humic acid (PPS-HA) was extracted from PPS with Na4P207 solution (Xing and Pignatello, 1997). The extract was centrifuged at 8600 g to remove undissolved material, and the HA was precipitated with HCl. The precipitate was then washed repeatedly with distilled water to remove chloride and 44 Table 3.1 Composition of sorbents Sorbent %C %H %N %ash Pahokee peat soil PPS 44.6 4.70 3.09 6.9 PPS-HA 51.8 3.22 3.25 5.6 PPS-humin 43.8 4.10 3.06 25.6 Michigan HA 51.8 4.17 2.96 0.9 freeze-dried. The insoluble solid remaining after extraction of PPS is referred to as PPS- humin. Cheshire soil (a Typic Dystrochrept) was a fine sandy loam from Lockwood Farm (Hamden, CT) with dry composition of 56% sand, 36% silt, 8% clay, and 1.4% organic carbon. Michigan humic acid was extracted from a Michigan peat (57% SOM, an Oxyaquic Hapludalf) with Na4P207 solution. The purification of the HA followed the same procedure detailed above. Michigan peat HA was used to prepare Ca- or Al- saturated HA. Briefly, the HA particles were saturated by four sequential equilibrations in 0.1 M CaCl2 or AICI3 solution, respectively. They were washed free of excess chloride by suspension in deionized water and centrifugation. Ca- or Al-saturated HAs were then freeze-dried and stored for subsequent use. Sorbates Atrazine (2-chloro-4-ethylamino-6-isopropylamino-s-triazine) and prometon (2,4- bis(isopropylamino)-6-methoxy-s-triazine) were from Crescent Chemical Co. 1,2- 45 Dichlorobenzene (1,2-DCB) and naphthalene were from Aldrich Chemical Company. [Ring-UL-''‘C]-atrazine, 1,2-DCB, and naphthalene were from Sigma Chemical Company. Sorption Isotherms Sorption experiments were conducted using batch equilibrations (Xing et al., 1996; Xing and Pignatello 1997) in 8-mL screw-cap vials (minimal headspace) with Teflon-lined septa. The solution was 0.01 M CaCl2 containing 200 mg/L HgCl2 as a biocide. In the case of the sorbent being Ca- or Al-saturated humic acid, the aqueous phase was 0.01 M CaCl2 or AICI3, respectively. The solution to solid ratio was adjusted to achieve 25%-85% uptake of the principle solute. Initial solution concentrations were up to half of the solute solubility in water. At 25°C, solubility of atrazine, 1,2-DCB and naphthalene are 33 pg/mL (Esser et al., 1975), 145 pg/mL and 33 pg/mL (Verschueren, 1996), respectively. Two controls without the sorbent were run at each initial concentration. The suspensions were shaken for 48 h (72 h for cation-saturated humic acids) in hematology mixers giving rocking-rotating motions. The vials were then centrifuged at 1000 g for 20 min, and the supernatants were sampled for liquid scintillation counting. The sorbed concentration was calculated by mass balance because sorption to vials was insignificant (Xing et al., 1996; Xing and Pignatello, 1997). The data for competitive sorption between atrazine and prometon were from a previous paper (Xing et al., 1996) and the procedure was described therein. 46 Data Analysis Sorption data were fitted with the common Freundlich model. The Freundlich parameters, Kf and N, were determined by linear regression (eq 3.9b) of log-transformed data. Linear fit of log-transformed data was justified over direct nonlinear curve fitting in this research for two reasons: (i) concentrations were spread evenly over the log scale, thus, nonlinear curve fitting would underestimate the importance of the low concentration data; and (ii) the relative uncertainty in the measurement was not greatly dependent on the concentration. Results and Discussions Examples of sorption isotherms are shown in Figure 3.1. The common Freundlich model fits well for all sorption data (r^ > 0.999, Table 3.2). Sorption is nonlinear in all systems, as indicated by values less than unity (Table 3.2). These data confirm further the heterogeneous nature of SOM and are consistent with other studies (Huang and Weber, 1998; Xing and Pignatello, 1997). Figure 3.2 shows the dependence of sorption energy on solute loading qe based on the relationship between qe and F^ defined in eq 3.10. The range of qe is determined by the range of equilibrium solution concentration, Ce, according to the Freundlich equation. In order to compare sorption of a single solute by different sorbents directly, loading is normalized to organic carbon content. Sorption site- energy initially decreases dramatically with increased loading, then decreases slowly. This confirms that there are 47 Table 3.2 Freundlich isotherm parameters logA:^ 7 System N r In Cheshire soil Atrazine 0.332 ±0.003 0.923 ± 0.004 1.00 Atrazine/prometon 0.154 ±0.007 0.987 ±0.010 0.999 1,2-DCB in PPS, whole 2.521 ±0.007 0.833 ±0.007 0.999 PPS-HA 2.146 ±0.007 0.896 ± 0.007 0.999 PPS-humin 2.626 ± 0.007 0.780 ±0.006 0.999 Naphthalene in PPS, whole 2.600 ± 0.006 0.790 ± 0.004 1.00 PPS-HA 2.363 ± 0.005 0.903 ± 0.004 0.999 PPS-humin 2.714 ±0.009 0.751 ±0.006 0.999 Naphthalene in Michigan HA 2.491 ±0.010 0.955 ± 0.008 0.999 Michigan Ca-HA 2.417 ±0.006 0.932 ± 0.005 0.999 Michigan Al-HA 2.396 ±0.011 0.868 ± 0.008 0.999 48 10000 o PPS-HA ° PPS-humin Figure 3.1 Sorption of naphthalene by different fractions of Pahokee peat soil (PPS). distributions of site energies and only a limited number of high energy sites in SOM which are filled first at low concentrations. The site-energy distributions with loading could easily explain the findings (Young and Weber, 1995; Spurlock, 1995), in which sorption heats change from significantly exothermic to less exothermic as the loading increases provided that the entropy change can be neglected compared to the free energy change during sorption. Sorption energy of atrazine in Cheshire soil is reduced by prometon as a co-soulte (Fig. 3.2A). This is because that prometon competes with atrazine for energic sites (Xing et al., 1996). 49 Sorption Energy E* (kJ/mol) 15 experimental systems.Theinsertsareplotsforlowconcentrationranges. (B) 1,2-DCBinPPS(...),PPS-HA(—), PPS-humin(--).(C)Naphthalenein (A) Atrazinealone(—),atrazinewith prometon(...)inCheshiresoil. HA (—),Ca-HAAl-HA(...). Figure 3.2Dependenceofsorptionenergyonsoluteloadingfordifferent PPS (...),PPS-HA(—),PPS-humin (--).(D)NaphthaleneinMichigan 0 4008001200 Solute Loadingq(^g/g) 50 Within the entire range of loading, sorption energy for PPS and its fractions is following the order: PPS-humin > PPS > PPS-HA (Figs. 3.2B, 3.2C). This is consistent with the order of sorption affinity constants (Kf values): PPS-humin > PPS > PPS-HA (Table 3.2). Sorption energies only have small changes when Michigan HA was saturated with Ca or A1 (Fig. 3.2D). This indicates that cations did not significantly affect sorption of solid HA particles. Figure 3.3 shows the site-energy distributions corresponding to the common Freundlich model. The frequency function decreases exponentially with increased site- energy. According to eq 8, the areas under the curves in Figure 3.3 represent the amount of sorption sites. Figure 3.3 A clearly shows that the number of sorption sites reduced after adding prometon as a co-solute. Moreover, from calculation we found that the percentage of site loss is higher at high energy region than at low energy region (e.g., at 16 kJ/mol, the loss is 42%, at 10 kJ/mol, 32%, at 5kJ/mol, 21%). This indicates that competitive effect is more significant at high energy sites. Loss of high energy sites leads to decrease of site heterogeneity (A^ value changed from 0.923 to 0.987, Table 3.2). This is coincident with the explanation by Xing et al. (1996) that the linearizing effect of prometon resulted from occupying and blocking the adsorption sites, while leaving the partition domain unaffected. Figures 3.3B and 3.3C show that the numbers of sorption sites for 1,2-DCB and naphthalene in different SOM fractions follow the order: PPS-humin > PPS > PPS-HA. This order is consistent with the prediction by the dual-mode model which postulates 51 F(E) (^g/g OC)(kJ/mol) Figure 3.3Approximatesite-energydistributionsderivedfromFreundlich Atrazine alone(—)andatrazinewithprometon(...)inCheshiresoil.(B). isotherm parameters.Theinsertsaretheplotsforhighenergyranges.(A). Michigan HA(—),Ca-HAand Al-HA(...). in PPS-huminPPS(...),andPPS-HA (—).(D).Naphthalenein 1,2-DCB inPPS-huminPPS(...), andPPS-HA(—).(C).Naphthalene E (kJ/mol) 52 more hole-filling sites in condensed domains. Humic acid is the readily extractable fraction of SOM, while humin is the insoluble organic matter remaining after extraction. It is reasonable to assume that the less condensed organic matter (fulvic and humic acids) is extracted preferentially. These less condensed organic fractions have lower population of adsorption sites (hole-filling sites). This is also manifested by the increase of N values of 1,2-DCB and naphthalene in the order: PPS-humin < PPS < PPS-HA (Table 3.2). The more condensed the SOM, the more nonlinear the sorption isotherm, and vice versa (Xing and Chen, 1999; Huang and Weber, 1997). It is interesting to note that although the numbers of sites reduce in cation- saturated Michigan HA at the low energy region, they increase at the high energy region (Fig. 3.3D). When HA was saturated by Al, the numbers of sites are reduced by 56% at sorption energy E* of 5 kJ/mol, while increased by 8% at £* of 20 kJ/mol. The crossing point is at FT of 17.5 kJ/mol. This tells us that new high energy sites may be created during the saturation of HA and the site-energy distributions become wider {N value changed from 0.955 for HA to 0.868 for Al-HA). The creation of high energy sites may be due to the conformational change of HA macromolecules by the cations. Hayes (1997) has pointed out that di- and trivalent cations may form inter- and intramolecular bridges between anionic sites of HA in solution. This causes the macromolecules to assume shrunken (or condensed) conformations. Multi-valent cations are known for their strong effects on dissolved HA structures and subsequent sorptive behaviors (Ghosh and Schnitzer, 1980; Traina et al., 1989; Schlautman and Morgan, 1993). 53 In summary, site-energy distribution is useful for explaining the experimental data of sorption. This analysis further demonstrates that SOM is heterogeneous in nature and lends additional support for nonlinear and competitive sorption. References Carter, M.C., J.E. Kilduff, and W.J. Weber, Jr. 1995. Site-energy distribution analysis of preloaded adsorbents. Environ. Sci. Technol. 29:1773-1780. Cerofolini, G.F. 1974. Localized adsorption on heterogeneous surfaces. Thin Solid Films. 23:129-152. Chiou, C.T. 1989. Theoretical considerations of the partition uptake of nonionic organic compounds by soil organic matter. In Reactions and Movement of Organic Chemicals in Soil. B.L. Sawhney, and K. Brown (eds.). Soil Science Society of America, Madison, WI. pp. 1-29. Derylo-Marczewska, A., and M. Jaroniec. 1984. Heterogeneity effects in single-solute adsorption from dilute solutions on solids. Chem. Scr. 24:239-246. Esser, H.O., G. Dupuis, E. Ebert, G. Marco, and C. Vogel. 1975. s-TRIAZINES, In Herbicides: Chemistry, degradation, and mode of action. P.C. Kearney, and D.D. Kaufman (eds.). Marcel Dekker, Inc. New York. pp. 129-208. Ghosh, K., and M. Schnitzer. 1980. Macromolecular structures of humic substances. Soil Sci. 129:266-276. Harris, L.B. 1968. Adsorption on a patchwise heterogeneous surface; mathematical analysis of the step-function to the local isotherm. Surface Sci. 10:129-145. Hayes, M.H.B. 1997. Emerging concepts of the compositions and structures of humic substances. In Humic substances in soils, peats and waters. Health and environmental aspects. M.H.B., Hayes, and W.S. Wilson (eds.). The Royal Society of Chemistry, Cambridge, UK. pp. 3-30. Huang, W., and W.J. Weber, Jr. 1997. A distributed reactivity model for sorption by soils and sediments. 10. Relationships between desorption, hysteresis, and the chemical characteristics of organic domains. Environ. Sci. Technol. 31:2562-2569. 54 Huang, W., and W.J. Weber, Jr. 1998. A distributed reactivity model for sorption by soils and sediments. 11. Slow concentration-dependent sorption rates. Environ. Sci. Technol. 32:3549-3555. Huang, W., T.M. Young, M.A. Schlautman, H. Yu, and W.J. Weber, Jr. 1997. A distributed reactivity model for sorption by soils and sediments. 9. General isotherm nonlinearity and applicability of the dual reactive domain model. Environ. Sci. Technol. 31:1703-1710. McGinley, P.M., L.E. Katz, and W.J. Weber, Jr. 1993. A distributed reactivity model for sorption by soils and sediments. 2. Multicomponent systems and competitive effects. Environ. Sci. Technol. 27:1524-1531. Schlautman, M.A., J.J. Morgan. 1993. Effects of aqueous chemistry on the binding of polycyclic aromatic hydrocarbons by dissolved humic materials. Environ. Sci. Technol. 27:961-969. Seidel, A., and P.S. Carl. 1989. The concentration dependence of surface diffusion for adsorption on energetically heterogeneous sorbents. Chemical Engineering Science. 44:189-194. Spurlock, F.C. 1995. Estimation of humic-based sorption enthalpies from nonlinear isotherm temperature dependence: theoretical development and application to substituted phenylureas. J. Environ. Qual. 24:42-49. Traina, S.J., D.A. Spontak, and T.J. Logan. 1989. Effects of cations on complexation of naphthalene by water-soluble organic carbon. J. Environ. Qual. 18:221-227. Verschueren, K. 1996. Handbook of environmental data on organic chemicals. Van Norstrand Reinhold, an International Thomson Publishing company. Weber, W.J., Jr., and W. Huang. 1996. A distributed reactivity model for sorption by soils and sediments. 4. Intraparticle heterogeneity and phase-distribution relationships under nonequilibrium conditions. Environ. Sci. Technol. 30:881-888. Weber, W.J., Jr., P.M. McGinley, L.E. Katz. 1992. A distributed reactivity model for sorption by soils and sediments. 1. Conceptual basis and equilibrium assessments. Environ. Sci. Technol. 26: 1955-1962. Xing, B. 1998. Reaction of toluene with soil organic matter. J. Environ. Sci. Health. 833:293-305. Xing, B., and J.J. Pignatello. 1997. Dual-mode sorption of low-polarity compounds in glassy poly(vinyl chloride) and soil organic matter. Environ. Sci. Technol. 31:792- 799. 55 Xing, B., and J.J. Pignatello. 1998. Competitive sorption between 1,3-dichlorobenzene or 2,4-dichlorophenol and natural aromatic acids in soil organic matter. Environ. Sci. Technol. 32:614-619. Xing, B., and Z. Chen. 1999. Spectroscopic evidence for condensed domains in soil organic matter. Soil Sci. 164:40-47. Xing, B., J.J. Pignatello, and B. Gigliotti. 1996. Competitive sorption between atrazine and other organic compounds in soils and model sorbents. Environ. Sci. Technol. 30:2432-2440. Young, T.M., and W.J. Weber, Jr. 1995. A distributed reactivity model for sorption by soils and sediments. 3. Effects of diagenetic processes on sorption energetics. Environ. Sci. Technol 29:92-97. 56 CHAPTER 4 SYNTHESIS Summary The overall objectives of this research are to provide better understanding of the interactions between organic chemicals and soil organic matter and to further confirm the applicability of the dual-mode model for sorption. Chapter 2 studied the effects of saturating cations on sorption/desorption by SOM for organic compounds varied with molecular size and polarity. Sorption driving forces for polar compounds may be different from those for nonpolar compounds as indicated by the different magnitude of hysteresis. Saturating cations may also play a role in determining sorption mechanisms. There was a good correlation between the magnitude of hysteresis and sorption nonlinearity. Chapter 3 performed a site-energy distribution analysis to study sorption mechanisms from a theoretical point of view. Typical systems have been chosen for this analysis. Sorption site-energy exhibited a distribution over the solid-phase concentration of the solute. Site-energy distribution analysis lends additional support for the dual-mode model. 57 Significance in the Field Although sorption of various classes of organic compounds by SOM has been intensively studied, desorption data were not sufficient at current stage. Sorption isotherms can be fitted with many models, among those the linear model and Freundlich model are mostly used. Fitting sorption data to a specific equation alone should not be used to determine sorption mechanisms because one set of data often fit well with several equations. Other evidence was needed to support the proposed mechanisms. Desorption experiments may serve this purpose. In addition, desorption data may provide direct information for successful clean-up practice. Sorption/desorption hysteresis is a common phenomenon in the solute-water-SOM systems. It will reduce the release of chemicals from SOM leaving a fraction of chemicals unrecoverable by conventional solvent washings. This research successfully explained the causes for hysteresis in sorption/desorption of organic compounds by SOM and constructed a good relationship between the magnitude of hysteresis and the sorption parameter, namely, Freundlich exponent. These findings have advanced our knowledge of the interactions between organic compounds and SOM. 58 Future Research 1. Broad classes of organic compounds, especially with different molecular size and polarity, need to be tested to examine the applicability of the dual-mode sorption mechanisms. 2. Sorbents other than HA, such as FA and humin, need to be examined to determine the correlations between the magnitude of hysteresis and Freundlich exponent, 3. In this study, equilibration periods for sorption and desorption are relatively short (3 days). Long-term (in weeks or months) sorption/desorption may behave differently, therefore, deserve to be investigated. 4. Thermodynamic parameters could help to check the feasibility of the proposed sorption mechanisms. Sorption experiments at different temperatures are needed. 59 APPENDIX A NMR SPECTRA OF HA 60 C-NMR CHARACTERIZATION OF HA < X! < d: < 13 X U C/2 (U B 03 c/i (U X • X on X U 'X X 00 X CJ (73 o Aiken, G.R., D.M. McKnight, R.L. Wershaw, and P. MacCarthy. 1985. An introduction to humic substances in soil, sediment, and water. In Humic substances in soil, sediment, and water: Geochemistry, isolation and characterization. G.R. Aiken, P. MacCarthy, R.I. Malcolm, and R.S. Swift (eds.). Wiley-Interscience, New York. pp. 1-9. 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Environ. Toxicol. Chem. 15,1282-1288. 65 Xing, B., and J.J. Pignatello. 1997. Dual-mode sorption of low-polarity compounds in glassy poly(vinyl chloride) and soil organic matter. Environ. Sci. Technol. 31:792- 799. Xing, B., and J.J. Pignatello. 1998. Competitive sorption between 1,3-dichlorobenzene or 2,4-dichlorophenol and natural aromatic acids in soil organic matter. Environ. Sci. Technol. 32:614-619. Xing, B., and Z. Chen. 1999. Spectroscopic evidence for condensed domains in soil organic matter. Soil Sci. 164:40-47. Xing, B., J.J. Pignatello, and B. Gigliotti. 1996. Competitive sorption between atrazine and other organic compounds in soils and model sorbents. Environ. Sci. Technol. 30:2432-2440. Young, T.M., and W.J. Weber, Jr. 1995. A distributed reactivity model for sorption by soils and sediments. 3. Effects of diagenetic processes on sorption energetics. Environ. Sci. Technol. 29:92-97. 66