Suthersan, Suthan S “Contaminant and Environmental Characteristics” Natural and Enhanced Remediation Systems Edited by Suthan S Suthersan Boca Raton: CRC Press LLC, 2001 ©2001 CRC Press LLC CHAPTER Contaminant and Environmental Characteristics CONTENTS 2.1 2.2 Introduction Contaminant Characteristics 2.2.1 Physical/Chemical Properties 2.2.1.1 Boiling Point 2.2.1.2 Vapor Pressure 2.2.1.3 Henry’s Law Constant 2.2.1.4 Octanol/Water Partition Coefficients 2.2.1.5 Solubility in Water 2.2.1.6 Hydrolysis 2.2.1.7 Photolytic Reactions in Surface Water 2.2.2 Biological Characteristics 2.2.2.1 Cometabolism 2.2.2.2 Kinetics of Biodegradation 2.3 Environmental Characteristics 2.3.1 Sorption Coefficient 2.3.1.1 Soil Sorption Coefficients 2.3.1.2 Factors Affecting Sorption Coefficients 2.3.2 Oxidation-Reduction Capacities of Aquifer Solids 2.3.2.1 pe and pH 2.3.2.2 REDOX Poise 2.3.2.3 REDOX Reaction References ©2001 CRC Press LLC Water is scientifically very different in comparison to other liquids With its rare and distinctive property of being denser as a liquid than as a solid, it is different Water is different in that it is the only chemical compound found naturally in solid, liquid, or gaseous states at ambient conditions Water is sometimes called the universal solvent This is a fitting name, especially when you consider that water is a powerful reagent, which is capable in time of dissolving everything on earth 2.1 INTRODUCTION The primary management goal during remediation of a contaminated site is to obtain closure, that is, to achieve a set of conditions that is considered environmentally acceptable and which will ensure that no future action will be required at the site A substantial ongoing national debate associated with site closure centers on the definition of “how clean is clean” for contaminated subsurface media The key issue in this debate is, “What concentration of residual contaminant in the subsurface, particularly adsorbed to the soil, is environmentally acceptable?” In this context, the term contaminant availability becomes an important concept; it refers to the rate and extent to which the chemical will be released from the subsurface into the environment and/or is bioavailable to ecological and human receptors The dissemination of a contaminant after its release into the environment is determined by its partition among the water, soil and sediment, and atmospheric phases, and its degradability via biotic and/or abiotic means These processes determine both the impact and the extent of its dissemination Within the context of overall site management, measurements of contaminant availability are not intended to replace other approaches, required regulatorily, to achieve site closure; rather, they are meant to broaden the range of options or tools available to environmental professionals This chapter will discuss the basis and parameters for the development of procedures and determination of partitioning, transport, and fate of various types of contaminants in the subsurface These parameters will also provide the basis for the development of the tools to determine contaminant availability and incorporate those estimations into a decision framework to define environmentally acceptable endpoints for the different media In addition, how these parameters and characteristics influence contaminant fate and transport and how they impact remediation system design are woven together in the discussions in subsequent chapters The reactions that contaminants undergo in the natural environment, such as sorption, desorption, precipitation, complexation, biodegradation, biotransformation, hydrolysis, oxidation-reduction, and dissolution, are critical in determining their fate and mobility in the subsurface environment Reaction time scales can vary from microseconds for many ion association reactions microseconds and milliseconds for some ion exchange and sorption reactions, to days, weeks, or months for some microbially catalyzed reactions, or years for many mineral solution and crystallization reactions Both transport and chemical reaction processes can affect the reaction rates in the subsurface environment Transport processes include: (1) transport in the solution phase, across a liquid film at the particle/liquid interface (film diffusion), and in ©2001 CRC Press LLC liquid-filled macropores, all of which are nonactivated diffusion processes and occur in mobile regions; (2) particle diffusion processes, which include diffusion of sorbate occluded in micropores (pore diffusion) and along pore-wall surfaces (surface diffusion) and diffusion processes in the bulk of the solid, all of which are activated diffusion processes (Figure 2.1).1 Pore and surface diffusion within the immediate region can be referred to as intra-aggregate (intraparticle) diffusion and diffusion in the solid can be called interparticle diffusion The actual chemical reaction at the surface, e.g., adsorption, is usually instantaneous The slowest of the chemical reaction and transport process is the ratelimiting reaction Film Solid (Soil Grain) Liquid (Groundwater) Figure 2.1 Transport in the Soil Solution (Macro Pores) Transport Across a Liquid Film at the Solid-Liquid Interface Transport in a Liquid-Filled Macropore Diffusion of a Sorbate at the Surface of the Solid Diffusion of a Sorbate Occluded in a Micropore Diffusion in the Bulk of the Solid Transport processes in solid-liquid soil reactions (adapted from Sparks, 1998) As an introduction to the various organic compounds which end up as contaminants once discharged into the environment, Table 2.1 gives the basic structure of the different compounds ©2001 CRC Press LLC Table 2.1 Some Common Functional Groups Example Functional Group General Formula General Name Formula IUPAC Name None CnH 2n+2 Alkane CH 3CH Common Name Ethane Ethane C C CnH 2n Alkene H 2C CH Ethene Ethylene C C CnH 2n-2 Alkyne HC CH Ethyne Acetylene Cl R Cl Chloride CH 2CH2Cl Chloroethane Ethyl chloride Br R Br Bromide CH3Br Bromomethane Methyl bromide OH R OH Alcohol CH 3CH2OH Ethanol Ethyl alcohol O R O Ether CH 3CH 2OCH2CH Ethoxyethane NH + NR X Amine RNH - C O H C R + R 4N X - C O CH3CH2CH Ketone R R C O DecyltrimethylAmmonium chloride DecyltrimethylAmmonium chloride Propanal Propionaldehyde Methyl ethyl ketone Acetic acid H OH CH CH3CH2C Carboxylic acid O OH Propylamine Ethanoic acid Aldehyde H - 2-Butanone CH3(CH2)9N(CH ) Cl R C O C + 3 Quaternary ammonium salt R C O 1-Aminopropane CH3CH2CH2NH2 Diethyl ether O O CH3 C OH (Continued) ©2001 CRC Press LLC Table 2.1 (Cont.) Example Functional Group O C OR' R C NH R C C Cl R C Amide NH O O O Acetamide Ethanoyl chloride Acetyl chloride Ethanoic anhydride Acetic anhydride Ethanenitrile Acetonitrile NH O Acid anhydride Acetic acid OC2H5 CH3 C O Common Name Ethanamide C O Acid chloride Cl Ethyl ethanoate CH C O IUPAC Name O CH OR' O O Formula Ester O O C General Name General Formula Cl O O C O C R C O C C R C NO R NO Nitro CH3 NO2 Nitromethane Nitromethane SH R SH Thiol CH3 SH Methanethiol Methyl mercaptan S R S R Thioether (sulfide) CH3 S CH Dimethyl thioether Dimethyl sulfide S R S S Disulfide CH S S Dimethyl disulfide Dimethyl disulfide Methanesulfonic acid OH R Methanesulfonic acid Dimethyl sulfoxide Dimethyl sulfoxide Dimethyl sulfone Dimethyl sulfone S N O S O O Sulfonic acid R S S O CH O OH O Sulfoxide Sulfone R O CH3 S R O R CH3 N CH S OH O O S R O O S CH3 C O C Nitrile N O S R CH3 O CH3 S CH O The italicized portion indicates the group A primary (1°) amine; there are also secondary (2°), R NH, and tertiary (3°), R amines N, Another name is propanamine ©2001 CRC Press LLC 2.2 2.2.1 CONTAMINANT CHARACTERISTICS Physical/Chemical Properties 2.2.1.1 Boiling Point The boiling point is defined as the temperature at which a liquid’s vapor pressure equals the pressure of the atmosphere on the liquid.2 If the pressure is exactly atmosphere (101,325 Pa), the temperature is referred to as “the normal boiling point.” Pure chemicals have a unique boiling point, and this fact can be used in some laboratory investigations to check on the identity and/or purity of a material Mixtures of two or more compounds have a boiling point range For organic compounds, boiling points range from –162 to over 700∞C, but for most chemicals of interest the boiling points are in the range of 300 to 600∞C.2 Having a value for a chemical’s boiling point, whether measured or estimated, is significant because it defines the uppermost temperatures at which the chemical can exist as a liquid Also, the boiling point itself serves as a rough indicator of volatility, with higher boiling points indicating lower volatility at ambient temperatures The boiling point is associated with a number of molecular properties and features Most important is molecular weight; boiling points generally increase with this parameter Next is the strength of the intermolecular bonding because boiling points increase with increasing bonding strength This bonding, in turn, is associated with processes and properties such as hydrogen bonding, dipole moments, and acid/base behavior 2.2.1.2 Vapor Pressure The vapor pressure of a chemical is the pressure its vapor exerts in equilibrium with its liquid or solid phase.2 Vapor pressure’s importance in environmental work results from its effects on the transport and partitioning of chemicals among the environmental media (air, water, and soil) The vapor pressure expresses and controls the chemical’s volatility The volatilization of a chemical from the water surface is determined by its Henry’s Law Constant, which can be estimated from the ratio of a chemical’s vapor pressure to its water solubility The volatilization of a chemical from the soil surface is determined largely by its vapor pressure, although this is tempered by its sorption to the soil matrix and its Henry’s Law Constant between the soil water content and air A substance’s vapor pressure determines whether it will occur as a free molecule in the vapor phase or will be associated with the solid phase For volatile substances that boil at or below 100∞C, the vapor pressure is likely to be known, but, for many high-boiling substances with low vapor pressure, the value may be unknown or poorly known An estimation procedure may be needed to help convert the known vapor pressure at the normal boiling point (i.e., atmosphere) to the vapor pressure at the lower temperatures of environmental importance For some of these high boiling compounds, the actual boiling point may also be unknown, since the substance may decompose before it boils ©2001 CRC Press LLC 2.2.1.3 Henry’s Law Constant Along with the octanol-water and octanol-air partition coefficients, the Henry’s Law Constant determines how a chemical substance will partition among the three primary media of accumulation in the environment, namely air, water, and organic matter present in soils, solids, and biota Volatile organic compounds (VOCs) with large values of Henry’s Law Constant evaporate appreciably from soils and water, and their fate and effects are controlled primarily by the rate of evaporation and the rate of subsequent atmospheric processes For such chemicals, an accurate value of this parameter KAW is essential Even a very low value of KAW for example, 0.001, can be significant and must be known accurately, because the volume of the accessible atmosphere is much larger than that of water and soils by at least a factor of 1000; thus even a low atmospheric concentration can represent a significant quantity of chemical Further, the rate of evaporation from soils and water is profoundly influenced by KAW because that process involves diffusion in water and air phases in series, or in parallel, and the relative concentrations which can be established in these phases control these diffusion rates.2,3 Accurate values of KAW are thus essential for any assessment of the behavior of existing chemicals or prediction of the likely behavior of new chemicals Air-water partitioning can be viewed as the determination of the solubility of a gas in water as a function of pressure, as first studied by William Henry in 1803 A plot of concentration or solubility of a chemical in water expressed as mole fraction x, vs partial pressure of the chemical in the gaseous phase P, is usually linear at low partial pressures, at least for chemicals which are not subject to significant dissociation or association in either phase This linearity is expressed as Henry’s Law The Henry’s Law Constant (H) which in modern SI units has dimensions of Pa/(mol fraction) For environmental purposes, it is more convenient to use concentration units in water CW of mol /m3 yielding H with dimensions of Pa m3/mol P (Pa) = H (Pa m3/mol) CW (mol/m3) (2.1) The partial pressure can be converted into a concentration in the air phase CA by invoking the ideal gas law: CA = n/V = P/RT (2.2) Where n is mols, V is volume (m3), R is the gas constant (8.314 Pa m3/mol K) and T is absolute temperature (K) CA = P/RT = (H/RT) CW = KAWCW (2.3) The dimensionless air-water partition coefficient KAW (which can be the ratio in units of mol/m3 or g/m3 or indeed any quantity/volume combination) is thus (H/RT) A plot of CA vs CW is thus usually linear with a slope of KAW as Figure 2.2 illustrates For organic chemicals which are sparingly soluble in water, these concentrations are limited on one axis by the water solubility and on the other by the ©2001 CRC Press LLC Concentration in Air CA (Vapor Pressure/RT) Slope = Kaw = H/RT [Solubility of Compound] Concentration in Water Cw Figure 2.2 Description of Henry’s Law Constant maximum achievable concentration in the air phase which corresponds to the vapor pressure, as Figure 2.2 shows To the right of or above the saturation limit, a separate organic phase is present Strictly speaking, this saturation vapor pressure is that of the organic phase saturated with water, not the pure organic phase.2,3 2.2.1.4 Octanol/Water Partition Coefficients The usefulness of the ratio of the concentration of a solute between water and octanol as a model for its transport between phases in a physical or biological system has long been recognized.2,4,5 It is expressed as POCT = CO /CW = KOW This ratio is essentially independent of concentration, and is usually given in logarithmic terms (log POCT or log KOW) The importance of bioconcentration in environmental hazard assessment and the utility of this hydrophobic parameter in its prediction led to an intense interest in the measurement of POCT and also its prediction from molecular structure (So many calculation methods have been published in the last five years that it is not possible to examine them all in detail.) 2.2.1.5 Solubility in Water Solubility in water is one of the most important physical chemical properties of a substance, having numerous applications to the prediction of its fate and its effects in the environment It is a direct measurement of hydrophobicity, i.e., the tendency of water to exclude the substance from solution It can be viewed as the maximum concentration which an aqueous solution will tolerate before the onset of phase separation ©2001 CRC Press LLC Substances which are readily soluble in water, such as lower molecular weight alcohols, will dissolve freely in water if accidentally spilled and will tend to remain in aqueous solution until degraded On the contrary, sparingly soluble substances dissolve more slowly and, when in solution, have a stronger tendency to partition out of aqueous solution into other phases They tend to have larger air–water partition coefficients or Henry’s Law Constants, and they tend to partition more into solid and biotic phases such as soils, sediments, and fish As a result, it is common to correlate partition coefficients from water to those media with solubility in water Solubility normally is measured by bringing an excess amount of a pure chemical phase into contact with water at a specified temperature, so that equilibrium is achieved and the aqueous phase concentration reaches a maximum value It is rare to encounter a single compound as the contaminant present in the groundwater at a contaminant site C* = C x i g i i i (2.4) where, Ci* Ci0 xi gi = = = = equilibrium solute concentration for component i in the mixture equilibrium solute concentration for component i as a pure compound mole fraction of compound i in the mixture activity coefficient of compound i in the mixture Possible equilibrium situations may exist, depending on the nature of the chemical phase, each of which requires separate theoretical treatment and leads to different equations for expressing solubility These equations form the basis of the correlations discussed later Single compound is an immiscible liquid (e.g., Benzene) C* = Co x g (2.5) In this case, C* is also C∞ Thus the product xg is 1.0 and x is 1/g Sparingly soluble substances act in such a way because the value of g is large.2 For example, at 25∞C benzene has a solubility in water of 1780 g/m3 or 22.8 mol/m.3 Since m3 of solution contains approximately 106/18 mol water (1m3 is 106 g and 18 g /mol is the molecular mass of water), the mole fraction x is 22.8/(106/18) or 0.00041 The activity coefficient g is thus 2440; i.e., a benzene molecule in aqueous solution behaves as if its concentration were 2440 times higher Substances such as polychlorinated biphenyls (PCBs) can have activity coefficients exceeding million Hydrophobicity thus is essentially an indication of the magnitude of g Some predictive methods focus on estimating g, from which solubility can be deduced ©2001 CRC Press LLC From this relationship it is apparent that the total organic carbon content of the aquifer matrix is less important for solutes with low octanol-water partitioning coefficients (Kow).47 Also apparent is the fact that the critical level of organic matter increases as the surface area of the mineralogic fraction of the aquifer matrix increases The surface area of the mineralogic component of the aquifer matrix is most strongly influenced by the amount of clay For compounds with low Kow values present in materials with a high clay content, sorption to mineral surfaces could be an important factor causing retardation of the chemical Several researchers have found that if the distribution coefficient is normalized relative to the aquifer matrix total organic carbon (TOC) content, much of the variation in observed Kd values between different soils is eliminated.49 Distribution coefficients normalized to total organic carbon content are expressed as Koc The following equation gives the expression relating Kd to Koc: K oc = Kd foc (2.19) where Koc Kd foc = soil sorption coefficient normalized for total organic carbon content = distribution coefficient = fraction of total organic carbon (mg organic carbon/mg soil) In areas with high clay concentrations and low TOC concentrations, the clay minerals become the dominant sorption sites Under these conditions, the use of Koc to compute Kd might result in underestimating the importance of sorption in retardation calculations, a source of error that will make retardation calculations based on the total organic carbon content of the aquifer matrix more conservative In fact, aquifers that have a high enough hydraulic conductivity to spread organic chemical contamination generally have a low clay content In these cases the contribution of sorption to mineral surfaces is generally trivial Sorption coefficients also have been expressed on an organic matter basis (Kom) by assuming that the organic matter content of a soil or sediment equals some factor, usually between 1.7 to 1.9, times its organic carbon content on a mass basis.47,50 Often 1.724 is used as this factor, implying that the carbon content of organic matter is 1/1.724 or 60% However, Koc is considered a more definite and less ambiguous measure than Kom.47 Assumptions inherent in the use of a Koc (or Kom) are that: sorption is exclusively to the organic component of the soil, all soil organic carbon has the same sorption capacity per unit mass, equilibrium is observed in the sorption–desorption process, and the sorption and desorption isotherms are identical.45 Both Koc and Kd have units of L/kg or cm3/g Numerous studies have been performed using the results of batch and column tests to determine if relationships exist that are capable of predicting the sorption characteristics of a chemical based on easily measured parameters The results of ©2001 CRC Press LLC these studies indicate that the amount of sorption is strongly dependent on the amount of organic carbon present in the aquifer matrix and the degree of hydrophobicity exhibited by the contaminant.47 These researchers observed that the distribution coefficient, Kd, was proportional to the organic carbon fraction of the aquifer times a proportionality constant This proportionality constant, Koc, is defined as given by Equation 2.19 Because it is normalized to organic carbon, values of Koc are dependent only on the properties of the compound (not on the type of soil) Values of Koc have been determined for a wide range of chemicals By knowing the value of Koc for a contaminant and the fraction of organic carbon present in the aquifer, the distribution coefficient can be estimated using the relationship Kd = Koc foc (2.20) The fraction of soil organic carbon must be determined from site-specific data Representative values of the fraction of organic carbon (foc ) in common sediments is available in the literature When predicting sorption of organic compounds, total organic carbon concentrations obtained from the most transmissive aquifer zone unaffected by contamination should be averaged and used for predictions This is because the majority of dissolved contaminant transport occurs in the most transmissive portions of the aquifer In addition, because the most transmissive aquifer zones generally have the lowest total organic carbon concentrations, the use of this value will give a conservative prediction of contaminant sorption and retardation Determination of the coefficient of retardation using sorption coefficients is described in Chapter 2.3.1.2 Factors Affecting Sorption Coefficients Many factors potentially can affect the distribution of a contaminant between an aqueous and solid phase These include environmental variables, such as temperature, ionic strength, dissolved organic matter concentration, and the presence of colloidal material, surfactants, and cosolvents In addition, factors related specifically to the experimental determination of sorption coefficients, such as sorbent and solid concentrations, equilibration time, and phase separation technique, can also be important A brief discussion of several of the more important factors affecting sorption coefficients follows Temperature: The effect of temperature on sorption equilibrium is a direct indication of the strength of the sorption process The weaker the interaction between sorbent and sorbate, the less the effect of temperature.47,50 While temperature can influence sorption, the strength and direction of the effect depends on the properties of the sorbent and sorbate and on the sorption mechanism Adsorption processes are generally exothermic, so the higher the temperature, the less the adsorption Hydrophobic sorption, however, has been shown to be relatively independent of temperature Other reviews also indicate that the influence of temperature on equilibrium sorption and have found that, in most cases, equilibrium sorption decreases with increasing temperature.47 ©2001 CRC Press LLC pH: For neutral chemicals, sorption coefficients usually are unaffected by pH However, for ionizable organic chemicals, sorption coefficients can be affected greatly, since pH affects not only the speciation but also the surface characteristics of natural sorbents Typically, for weak acids the free acid form (HA) is more strongly sorbed than the anionic form (A–) For example, pentacholorophenol (PCP) sorption decreased with increasing pH over the entire pH range tested (2 to 12) For weak bases the cationic form dominates at low pH and is more highly sorbed than the free base.44 Ionic Strength: Salts can affect sorption of organic compounds by displacing cations from the soil ion exchange matrix, by changing the activity of the sorbate in solution, and by changing the charge density associated with the soil sorption surface Salt effects are most important for basic sorbates in the cation state, where an increase in salinity can significantly lower the sorption coefficient Salt effects are least important for neutral compounds, which may show either increases or decreases in sorption as salinity increases.44 Dissolved or Colloidal Organic Matter: The presence of dissolved or colloidal organic matter has been shown to influence sorption depending on the nature of the chemical and the organic matter Some compounds were found to be associated extensively with the dissolved organic matter; sorption by soil decreased significantly in the presence of dissolved organic matter Some have characterized several size fractions of water soluble organic carbon and found that the effect of dissolved organic matter on the sorption of pyrene may be limited, but the presence of colloidal organic matter suspended in the soil solution may have significant impact on the sorption of pyrene.44,51 Cosolvents: The effect of nonpolar cosolutes (trichloroethylene, toluene), polar cosolutes (1-octanol, chlorobenzene, nitrobenzene, o-cresol) and polar cosolvents (methanol and dimethyl sulfoxide) on sorption of several polycyclic aromatic hydrocarbons (PAHs) has been investigated.44,52 The nonpolar cosolutes did not significantly influence PAH sorption, while the polar cosolutes (nitrobenzene, o-cresol), having sufficiently high aqueous solubilities, caused a significant decrease in PAH sorption Miscible organic solvents, such as methanol and ethanol, have been shown to increase solubility of hydrophobic organics and to decrease sorption This is presumably the result of reducing the activity coefficient of the sorbate chemical in the aqueous phase, and competition for sorbing sites Competitive Sorption: At concentrations normally encountered in environmental situations, sorption often has been observed to be relatively noncompetitive For example, it was found that there is no competition in the sorption of binary solutes m-dichlorobenzene and 1,2,4-trichlorobenzene and between parthion and lindane.53 The sorption of methyl and dimethyl naphthalene, individually and as components of JP-8 and synthetic jet fuel mixture, on two sediments and montomorillonite clay in water was measured.54 The sorption coefficients of the naphthalenes generally varied by less than a factor of two However, there are reports of competitive sorption taking place that is thought to be the result of site-specific sorption occurring in soil organic matter ©2001 CRC Press LLC Organic Matter Type and Origin: While the constancy of Koc values suggests a uniformity of organic matter with regard to sorption behavior, it is becoming increasingly apparent that organic matter type can be an important sorption variable for some sorbent/sorbate combinations For example, it was found that the sorption of naproamide, a nonionic herbicide, was greater in the sediment than in soils, even on an organic carbon basis.44 The increased sorption in sediment was attributed to the fact that soils contained a higher percentage of cellulose and hemicellulose material, whereas the sediments contain a higher lipid-like fraction Kinetic Considerations: Sorption generally is regarded as a rapid process and, in many laboratory sorption experiments, equilibrium often is observed within several minutes or hours An equilibration time of 24 hours often is used for convenience True sorption equilibrium under natural conditions, however, may require weeks to months to achieve depending on the chemical and environmental solid of interest In many instances, an early period of rapid and extensive sorption, followed by a long slow period, is observed Experimental determination of sorption coefficients requires preliminary kinetic experiments to determine the time to reach equilibrium Two processes govern rate-limited or nonequilibrium sorption: transport of the substance to the sorption sites and the sorption process itself.44,50 Transportrelated nonequilibrium typically results from the existence of a heterogeneous flow domain Sorption-related nonequilibrium, caused by rate-limited interactions between the sorbate and sorbent, may be the result of chemical nonequilibrium (i.e., chemisorption) or diffusive mass transfer limitations (i.e., diffusion of solute within pores of microporous particles or molecular diffusion into macromolecular organic matter) Sorption kinetics are likely to be environmentally important in short contact situations such as sediment resuspension, soil erosion, and infiltrating ground water.44 In general, adsorption processes tend to be rapid and nearly instantaneous, whereas nonsurface sorption tends to be slower For neutral organic chemicals, the more hydrophobic the compound, the larger the sorption coefficient, and the longer it takes to reach equilibrium between the solid and aqueous phases This is because the sorbent must remove a chemical from a larger volume of water Generally, sorption estimates are based on equilibrium conditions only; however, incorporation of kinetic considerations into sorption estimation techniques is likely to be an important area of future work For example, the assumption of equilibrium sorption in dynamic field systems may result in calculating too much pesticide in the sorbed state Ionizability: For neutral organic compounds, in soils having a low clay/organic carbon ratio, sorption coefficients tend to increase as the hydrophobicity of the compound increases Aqueous solubility or octanol/water partition coefficients often are used as indicators of a compound’s hydrophobicity An increase on polarity, number of functional groups, and ionic nature of the chemical will increase the number of potential sorption mechanisms for a given chemical For ionizable compounds, pKa is of particular importance because it determines the dominant form of a chemical at the specific environmental pH ©2001 CRC Press LLC The entropy change is largely due to the destruction of the highly structured water shell surrounding the solvated organic The term “partitioning” was used to denote an uptake in which the sorbed organic chemical permeates the network of an organic medium by forces common to the solution, analogous to the extraction of an organic compound from water with an organic liquid By either description, hydrophobic sorption or partitioning should increase as compounds become less water soluble or more hydrophobic Additional characteristics typically associated with hydrophobic sorption or partitioning include sorption isotherms that are linear over a relatively wide range of concentrations, and sorption coefficients that are not strongly temperature dependent, and lack a competition between sorbates.44,53 2.3.2 Oxidation-Reduction Capacities of Aquifer Solids There has been considerable research activity focused on the characterization of REDOX-potential or intensity (Eh) conditions in groundwater systems defined as the REDOX activity of dissolved chemical species Early observations of significant Eh trends along groundwater flow paths led to hypotheses of successive REDOX zones characterized by the activity of specific thermodynamically favored electron acceptors These REDOX zones may be classified as oxic (i.e., detectable dissolved O2), suboxic or postoxic (i.e., no detectable O2 or sulfide, detectable Fe2+), and reducing (i.e., detectable Fe2+ and sulfide, no detectable O2).1 Further investigations correctly postulated that oxidation-reduction processes were mediated by natural microbial populations that catalyze electron-transfer reactions More recent work noted considerable temporal and spatial variability in measured subsurface REDOX conditions and that the succession of electron acceptors under oxic, suboxic, or reducing conditions was not strictly predictable by either chemical equilibrium calculations or platinum electrode measurements 2.3.2.1 pe and pH A pH is the negative log (p for power) of proton (H+) activity and pe, its energy or work analog, is the negative log of the electron potential An electron is not a full-fledged analog of a proton Together, two equal but opposite charges make up a hydrogen atom, but that is about the extent of the equality between an electron and a proton Without its proton, an electron is no longer an analog of H+, and it no longer has any claim to being part of a hydrogen atom An electron does not bounce about by itself in the manner of an H+, and therefore it is probably not correct to try to characterize its “activity.” It always is either attached to an atom or radical or in the process of being transferred from one to another A proton is a cation It can replace or be replaced by other cations and it is as good as any other cation when it comes to balancing a chemical equation Electrons receive no recognition in balanced chemical equations because the donated and accepted electrons must always cancel one another on opposite sides of an equation ©2001 CRC Press LLC Electrons not have anion status They cannot trade places with other negatively charged species Usually we see release of H+ when metals are oxidized and consumption of H+ with their reduction Oxidation is furthered in a subsurface environment where protons and electrons are deficient; that is, where acidity and levels of easily degraded (labile) electron donors are low But there must be a ready supply of available electron acceptors Reduction is favored by surpluses of both protons and electrons This means that low pH and high availability of organic substances will promote reduction in soil Reduction of Fe or Mn oxides, or of nitrate, uses up H+, thereby increasing pH of the soil and, theoretically, lowering the pe Oxidation of Fe, Mn, or nitrate lowers the pH (measurable) and raises the pe (not measurable in most soils) Measuring changes in concentrations of REDOX species is more reliable for predicting these things in the subsurface than is an attempted measurement of pe with the platinum electrode The farther apart the electrons, the more proportional work required to bring them together and the higher the respective pe A low pe system has a surplus of electrons and, therefore, a big tendency to lose some of them and become oxidized A high pe system is hungry for electrons As deficient electrons are replenished, the tendency for reduction to occur will increase If we substitute pe and pH for their defined equivalents in a generic REDOX half-reaction in which activities of oxidized and reduced species are equal, we see that the (pe + pH) sum is equivalent to the equilibrium constant of the half-reaction: Oxidized species + e– + H+ = reduced species (2.21) log K = log red – log ox – log e– – log H+ (2.22) log K = pe = pH (2.23) If indeed their sum is constant, then, thermodynamically, pe and pH are on opposite ends of a seesaw If behavior follows thermodynamic theory, when one goes up, the other will come down, like any sound seesaw This sum is referred to as the REDOX parameter because, if a soil is at internal equilibrium, the (pe + pH) represents the sums of all of the REDOX equilibrium constants in the soil.1 2.3.2.2 REDOX Poise In the natural environment REDOX seesaws are not so simple This seesaw-like behavior reflects the interaction between source/sink quantities and electron/proton intensities If we add reducing reagents or reduced substances such as Fe(II), or Mn(II) or Cr(III) to a soil poised so that its easily reduced substances are in balance with its easily oxidized substances, some of the added reduced species will be quickly oxidized On the other hand, adding Fe(III), Mn(IV) or Cr(VI) will result in immediate reduction of a portion of the added oxidants There appears to be a tendency for a soil, if disturbed, to maintain a REDOX balance, that is, poise, by donating ©2001 CRC Press LLC electrons to surplus electron acceptors or by accepting electrons from surplus electron donors.1 A soil kept near field capacity moisture with occasional mixing, double bagged inside a thin polyethylene bag for several months at 15 to 25ºC, will be close to internal equilibrium If this metastable equilibrium is disturbed by adding an easily oxidized substance to it, e.g., glucose, the (pe +pH) of the overall system will tend to remain fairly constant as the disturbed soil system moves back toward a new metastable equilibrium In this instance, the pe will tend to go down, and to the extent that it does, the pH will tend to rise.1 By adding increments of Cr3+ and HCrO4– , respectively, to separate subsamples of the same soil and then determining the amount of Cr reduced [loss of Cr(VI)] and the amount oxidized [gain of Cr(VI)], it is possible to find a point of poise or buffered REDOX region, where the electron donating and electron accepting tendencies cross There the REDOX seesaw is balanced at dead-level.1 2.3.2.3 REDOX Reactions REDOX is one of those catchy phrases invented by someone unhampered by commitment to the use of scientifically correct terminology The name is reversed (RED-OX, instead of OX-RED) for the sake of easy pronunciation The RED stands for reduction and it signifies gain of electrons by a chemical species called electron acceptors; the OX connotes oxidation, or electron loss by a chemical species called electron donors Oxidation-reduction (REDOX) reactions, along with hydrolysis and acid-base reactions, account for the vast majority of chemical reactions that occur in aquatic environmental systems (soils, sediments, aquifers, rivers, lakes, and many remediation operations) This section provides a survey of the environmental and substrate characteristics that govern REDOX transformations in aquatic systems The distinction between biotic and abiotic processes is a particularly important issue in defining the scope of this section Living organisms are responsible for creating the conditions that determine the REDOX chemistry of most aquatic environmental systems So, in this sense, most REDOX reactions in natural systems ultimately are driven by biological activity Once environmental conditions are established, however, many important REDOX reactions proceed without further mediation by organisms These reactions are considered to be abiotic when it is no longer practical (or possible) to link them to any particular biological activity Assigning Oxidation States: REDOX reactions involve oxidation and reduction; they occur by the exchange of electrons between reacting chemical species.2,55 Electrons (or electron density) are lost (or donated) in oxidation and gained (or accepted) in reduction An oxidizing agent (or oxidant) that accepts electrons (and is thereby reduced) causes oxidation of a species Similarly, reduction results from reaction with a reducing agent (or reductant) that donates electrons (and is oxidized) To interpret REDOX reactions in terms of electron exchange, one must account for electrons in the various reacting species Various textbooks provide simple rules, such as the following, for assigning oxidation states for inorganic REDOX couples:2,55 â2001 CRC Press LLC ã ã ã ã ã For free elements, each atom is assigned oxidation number Monoatomic ions have an oxidation number equal to the charge of the ion Oxygen, in most compounds, has the oxidation number –2 Hydrogen, in most compounds, has the oxidation number +1 Halogens, in most environmentally relevant compounds, have the oxidation number –1 These rules, however, are not easily applied to organic REDOX reactions, and this difficulty has led to a steady stream of alternative concepts for assigning oxidation states For present purposes, familiarity with a method for assigning oxidation states to organic molecules is sufficient This method reflects the qualitative observations from which the historical concepts of oxidation and reduction originated: oxidation is the gain of oxygen (O), chlorine (Cl) or double bonds, and/or the loss of H; reduction is the gain of H, saturation of double bonds, and/or loss of O or Cl Thus, for example, mineralization of any hydrocarbon to CO2 and H2O involves oxidation, and dechlorination of any chlorinated compound to hydrocarbon products involves reduction Oxidations: Organic chemicals that are susceptible to oxidation and are of concern from the perspective of contamination and environmental degradation include aliphatic and aromatic hydrocarbons, alcohols, aldehydes, and ketones, phenols, polyphenols, sulfides (thiols), sulfoxides, nitriles, amines, diamines, nitrogen and sulfur hetercyclic compounds, mono- and di-chlorinated aliphatics and many others Equations below show example half-reactions for oxidation of some of these chemical groups Alkanes to alcohols R – H + H2O Ỉ R – OH + 2H+ + 2e– (loss of H+ and e–) (2.24) Alcohols to aldehydes R CH2 OH Ỉ RCHO + 2H+ + 2e– (loss of H+ and e–) (2.25) Aldehydes to acids RCHO + H2O Ỉ RCOOH + 2H+ + 2e– (loss of H+ and e–) (2.26) Reductions: Most interest in reductive transformations of environmental chemicals involves dechlorination of chlorinated aliphatic and aromatic compounds and the reduction of nitroaromatic compounds Other examples of reductive transformations that may occur abiotically in the environment include reduction of azo compounds, quinines, disulfides, and sulfoxides An example of a half-reaction is described by the equation: ©2001 CRC Press LLC Reductive dechlorination R – Cl + H+ + 2e– Ỉ R – H + Cl– (gain of H+, e– and loss of Cl–) (2.27) Dechlorination can occur by several reductive pathways The simplest results in replacement of a C-bonded halogen atom with a hydrogen and is known as hydrogenolysis or reductive dechlorination The process is illustrated for trichloroethene, TCE, in Figure 2.12, where complete dechlorination by this pathway requires multiple hydrogenolysis steps The relative rate of each step is a critical concern because the steps tend to become slower with each dechlorination (and DCE and VC are at least as hazardous as TCE if not more so than with VC) Aryl halogens, such as those in the pesticide chlophyrifos, also are subject to hydrogenolysis, but this reaction rarely occurs abiotically One notable exception is the rapid abiotic dechlorination of polychlorinated biphenyls (PCBs) by zero-valent iron with catalysis by Pd.2,55 H CI C +H+ +2e-CI- C CI H CI C CI TCE Figure 2.12 H H +H+ +2e-CI- C CI CI C cis-1,2-DCE +H+ +2e-CI- C H H H H C C H VC H Ethene Reductive dechlorination or hydrogenolysis of TCE The other major dechlorination pathway involves elimination of two chlorines, leaving behind a pair of electrons that usually goes to form a carbon-carbon double bond Where the pathway involves halogens on adjacent carbons, it is known as vicinal dehalogenation or reductive b-elimination The major pathway for reductive transformation of lindane involves vicinal dehalogenation, which can proceed by steps all the way to benzene (Figure 2.13).2,55 Recently, data have shown that this pathway not only can convert alkanes to alkenes, but also can produce alkynes from dihaloalkenes (see Equation 2.28) CI CI CI CI CI CI CI +2e-2CI- CI CI CI CI +2e-2CI- CI Lindane Figure 2.13 Vicinal dechlorination or reductive-elimination of lindane ©2001 CRC Press LLC +2e-2CI- Benzene Vicinal Dehalogenation Cl – R – R1 – Cl + 2e– Ỉ R = R1 + 2Cl– (formation of double bond) (2.28) The contaminant REDOX reactions just summarized only occur when coupled with suitable half-reactions involving oxidants or reductants from the environment In a particular environmental system, these REDOX agents collectively determine the nature, rate, and extent of contaminant transformation Under favorable circumstances, the dominant REDOX agent(s) can be identified and quantified, thereby providing a rigorous basis for estimating the potential for, and rate of, transformation by abiotic REDOX reactions.2,55 Such specificity is often possible with systems engineered for contaminant remediation However, natural systems frequently involve complex mixtures of REDOXactive substances that cannot be characterized readily The characterization of REDOX conditions in complex environmental media is a long-standing challenge to environmental scientists that continues to be an active area of research The remainder of this section summarizes what is currently known about the identity of oxidants and reductants relevant to environmental systems, in order to provide a basis for estimating rates of contaminant transformations by specific pathways With respect to natural reductants, however, a great deal remains to be learned, so substantial developments can be expected as new research in this area becomes available Oxidants: The best opportunities for predicting REDOX transformations come from engineered systems where a known oxidant is added to achieve contaminant remediation Well-documented examples include the use of ozone and chlorine in drinking water treatment In natural systems, important oxidants are oxides of iron and manganese, as well as molecular oxygen and various photooxidants In engineered remediation systems oxidants used include potassium permanganate, ozone and hydrogen peroxide.1 The presence of molecular oxygen, O2 is used widely as the defining characteristic of oxidizing environments because the overwhelming supply of molecular oxygen makes it the ultimate source of oxidizing equivalents However, O2 in its thermodynamic ground-state (3O2) is a rather poor oxidizing agent and it is not usually the oxidant directly responsible for oxidative transformations of contaminants Instead, activated oxygen species may be involved where they are formed by the action of light on natural organic matter (NOM), peroxides, or various inorganic catalysts Activated oxygen species include singlet oxygen (1O2), protonated super· oxide (HO2 ) hydrogen peroxide and hydroperoxide anion (H2O2/HO2– ), hydroxyl radical (OH•), and ozone (O3).1,2,55 O2 + e– + H+ = H2O˚ (protonated superoxide) (2.29) O2 + 2e– + 2H+ = H2O2 (hydrogen peroxide) (2.30) O2 + 3e– + 3H+ = H2O + HO˚ (hydroxyl free radical) (2.31) ©2001 CRC Press LLC O2 + 4e– + 4H+ = 2H2O (water) (2.32) Equations 2.29–2.32 are half-reactions showing reduction of O2 by single electron additions Thus, superoxide and hydroxyl, produced by one and three odd electron additions, are free radicals; whereas peroxide and water, with two and four electrons added, respectively, are not Restricting conditions of interaction between the availabilities of soil O2 and electron donors, for example, at the interface between oxygenated water and anaerobic soil in a wetland, tends to favor transfers of electrons in single steps, and thus such interfaces are likely to be sites for free radical formation Free radical mechanisms appear to explain why kinetically slow and seemingly unlikely REDOX transformations often occur readily at interfaces Oxygen free radicals are much more reactive than O2 itself, and both superoxide and the hydroxyl free radical are especially reactive with H2O2, each one capable of being quickly transformed into the other Aside from oxygen and the activated oxygen species, there are several other oxidants that cause abiotic oxidation reactions involving environmental contaminants In engineered systems, these include chlorine, chlorine dioxide, permanganate and ferrate At highly contaminated sites, anthropogenic oxidants such as chromate, arsenate, and selenate may react with co-contaminants such as phenols In natural anoxic environments, the major alternative oxidants are Fe(III) and manganese (IV) oxides and hydroxides Both are common in natural systems as crystalline or amorphous particles or coatings on other particles In the absence of photocatalysis, however, iron and manganese oxides are weak oxidants As a result, they appear to react at significant rates only with phenols and anilines In the dissolved phase, few alternative abiotic oxidants are available in the neutral environment Nitrate, sulfate, and other terminal electron acceptors used by anaerobic microorganisms are thermodynamically capable of oxidizing some organic contaminants, but it appears that these reactions almost always require microbial mediation Reductants: Abiotic environmental reductants are not well characterized as the oxidants because, until recently, there were fewer remediation applications of reductants, and natural reducing environments are characterized by especially complex biogeochemistry The most familiar natural reductants are sulfide (present primarily as HS– and H2S), Fe (II) and Mn (II), and natural organic matter (NOM) The transformation of contaminants by sulfur species in anaerobic environments can involve both reduction and nucleophilic substitution pathways These processes have been studied extensively, but the complex speciation of sulfur makes routine predictions regarding these reactions difficult.1,2,55 A similar situation applies for reduced forms of iron As with oxidations, some of the best opportunities for reliably estimating rates of redox transformations are afforded by engineered systems where a reductant of known composition and quantity is added to achieve contaminant remediation In addition to zero-valent iron, other methods for chemical reduction of contaminants involve dithionite and electrolysis (where, in effect, electrons are added directly).1,2,55 The role of natural organic reductants in environmental systems is even more difficult to characterize than the roles of sulfur and iron because most natural organic matter is of indeterminate composition There are two general categories of NOM: ©2001 CRC Press LLC high molecular weight organic materials such as humic and fulvic acid, and low molecular weight compounds such as acids, alcohols, etc Specific examples of the latter include glycolate, citrate, pyruvate, oxalate, and ascorbate These types of compounds have been studied extensively for their role in global cycling of carbon, but very little work has been done on whether they act as specific reductants of organic contaminants.1,2,55 In contrast, the possibility that high molecular weight NOM acts as a reductant in environmental systems is widely acknowledged Although most evidence for this involves the reduction of metal ions, several studies have shown that the process extends to various organic contaminants Presumably, the reducing potential of NOM is due to specific moieties such as complex metals or conjugated polyphenols Often, REDOX reactions involving these moieties are reversible, which means that NOM may serve as a mediator of REDOX reactions rather than being just an electron donor (or acceptor).1,2,55 In the recent past, the addition of labile electron donors such as molasses, lactate, and methanol is gaining ground to facilitate enhanced reductive dechlorination of chlorinated aliphatic and aromatic compounds This technology is discussed in detail in Chapter Demonstrating that a REDOX transformation of a contaminant involves mediated electron transfer requires meeting several criteria: 1) the overall reaction must be energetically favorable, 2) the mediator must have a reduction potential that lies between the bulk donor and the terminal acceptor so that both steps in the electron transfer chain will be energetically favorable, and 3) both steps in the mediated reaction must be kinetically fast relative to the direct reaction between bulk donor and terminal acceptor Most evidence for involvement of mediators in reduction of contaminants comes from studies with model systems, because natural reducing media (such as anaerobic sediments) consist of more REDOX couples than can be characterized readily Although this is an active area of research, a variety of likely mediator half-reactions can be identified REFERENCES Sparks, D L., Soil Physical Chemistry, CRC Press, Boca Raton, FL, 1998 Boethling, R S and D MacKay, Handbook of Property Estimation Methods for Chemicals, Lewis Publishers, Boca Raton, FL, 2000 MacKay, D., W Y Shiu, and K C Ma, Henry Law Constant, in Handbook of Property Estimation Methods for Chemicals, Boethling, R S and D MacKay, Eds., Lewis Publishers, Boca Raton, FL, 2000 Leo, A et al., Partition coefficients and their uses, Chem Rev., 71, 525–616, 1971 Leo, A J., Hydrophobicity, the underlying property in most biochemical events, Environmental Health Chemistry, McKinney, J., Ed., Ann Arbor Science, Ann Arbor, MI, 1981, 323–336 Kenage, E., Determination of bioconcentration potential, Residue Rev., 44, 73–113, 1996 Neely, W B et al., Partition coefficients to measure bioaccumulation potential of organic chemical in fish, Environ Sci Technol., 8, 1113–1115, 1974 ©2001 CRC Press LLC Lyman, W J., W F Reehl, and D H Rosenblatt, Handbook of Chemical Property Estimation Methods, McGraw-Hill, New York, 1982 Wolfe, N L., and P M Jeffers, Hydrolysis, in Handbook of Property Estimation Methods for Chemicals, Boethling, R.S and D MacKay, Eds., Lewis Publishers, Boca Raton, FL, 2000 10 Wolfe, N L., Organophosphate and organophosphorothioate esters: application of linear free energy relationships to estimate hydrolysis rate constants for use in environmental fate assessment, Chemosphere, 9, 571–579, 1980 11 Mabey, W R and T Mill, Critical review of hydrolysis of organic compounds in water under environmental conditions, J Phys Chem Ref Data, 7, 383–415, 1978 12 Jeffers, P M et al., Homogeneous hydrolysis rate constants for selected methanes, ethanes, ethenes and propanes, Environ Sci Technol., 23, 965–969, 1989 13 Mill, T., Photoreactions in surface waters, in Handbook of Property Estimation Methods for Chemicals, Boethling, R S and D MacKay, Eds., Lewis Publishers, Boca Raton, FL, 2000 14 Larson, R A., L L Hunt, and D W Blankenship, Formation of toxic products from a No fuel oil by photooxidation, Environ Sci Technol., 11, 492–496, 1977 15 Atkinson, R J., A structure-activity relationship for the estimation of rate constants for the gas phase reactions of OH radicals with organic compounds, Int J Chem Kinetics, 19, 799–828, 1987 16 Hoag, W R and T Mill, Survey of sunlight-produced transient reactants in surface waters, Proceedings of a workshop on effects of solar ultraviolet radiatiaon on geochemical dynamics, Woods Hole, MA, 1989 16a Atkinson, R., Atmospheric Oxidation, in Handbook of Property Estimation Methods for Chemicals, Boethling, R S and D MacKay, Eds., Lewis Publishers, Boca Raton, FL, 2000 17 Mopper, K and X Zhou, Hydroxyl radical photoproduction in the sea and its potential impact on marine processes, Science, 250, 661–664, 1990 18 Howard, P H., Biodegradation, in Handbook of Property Estimation Methods for Chemicals, Boethling, R S and D MacKay, Eds., Lewis Publishers, Boca Raton, FL, 2000 19 Alexander, M., Biodegradation and Bioremediation, Academic Press, New York, 1999 20 Spain, J C and P A Van Weld, Adaptation of natural microbial communities to degradation of xenobiotic compounds: effects of concentration, exposure time, inoculum, and chemical structure, Appl Environ Microbiol., 45, 428–435, 1983 21 Howard, P H and S Banerjee, Interpreting results from biodegradability test of chemicals in water and soil, Environ Toxicol Chem., 3, 551–562, 1984 22 Alexander, M., Biodegradation of organic chemicals, Environ Sci Technol., 19, 106–111, 1985 23 Taylor, B F et al., Arch Microbio., 122, 301–306, 1979 24 Oldenhuis, R et al., Appl Environ Microbiol., 55, 2816–2819, 1989 25 Nelson, M J K et al., Appl Environ Microbiol., 54, 604–606, 1988 26 Li, S and L P Wackett, Biochem Biophy Res Commun., 185, 443–451, 1992 27 Rebertson, J B et al., J Appl Environ Microbiol., 58, 2643–2648, 1992 28 Delgado, A et al., J Appl Environ Microbiol., 58, 415–417, 1992 29 Shields, M S et al., J Appl Environ Microbiol., 57, 1935–1941, 1991 30 Wackett, L P et al., J Appl Environ Microbiol., 55, 2960–2964, 1989 31 Hyman, M R et al., J Appl Environ Microbiol., 60, 3033–3035, 1994 ©2001 CRC Press LLC 32 Van Beilen, J B., J Kingma, and B Witholt, Eng Microb Technol., 16, 904–911, 1994 33 Lee, K and D T Gibson, J Appl Environ Microbiol., 62, 3101–3106, 1996 34 Hernandez, B S., J J Arensdorf, and D D Focht, Biodegradation, 6, 75–82, 1995 35 Ladd, T I et al., Heterotropic activity and biodegradation of labile and refractory compounds in groundwater and stream microbial population, Appl Environ Microbiol., 44, 321–329, 1982 36 Neilson, A H., Organic Chemicals, Lewis Publishers, Boca Raton, FL, 1999 37 Alexander, M., Biodegradataion of chemicals of environmental concern, Science, 211, 132–138, 1981 38 Klopman, G et al., Computer-automated predictions of aerobic biodegradation of chemicals, Environ Toxicol Chem., 14, 395–403, 1995 39 Punch, W F et al., Bess, a computerized system for predicting the biodegradation potential of new and existing chemicals, 7th Int Workshop on QSARS in Env Sci., June 24-28, Elsinore, Denmark, 1996 40 Alexander, M., Nonbiodegradable and other recalcitrant molecules, Biotechnol Bioeng., 15, 611–647, 1973 41 Howard, P H et al., Review and Evaluation of Available Techniques for Determining Persistence and Routes of Degradation of Chemical Substances in the Environment, EPA-560/5-75-006, U.S NTIS PB 243825, 1975 42 Simkins, S and M Alexander, Models for mineralization kinetics with the variables of substrate concentration and population density, Appl Environ Microbiol., 47, 1299–1306, 1984 43 Schmidt, S K., S Simkins, and M Alexander, Models for the kinetics of biodegradation of organic compounds not supporting growth, Appl Environ Microbiol., 50, 323–331, 1985 44 Doucette, W J., Soil and Sediment Sorption Coefficients, in Handbook of Property Estimation Methods for Chemicals, Boethling, R S and D MacKay, Eds., Lewis Publishers, Boca Raton, FL, 2000 45 Green, R E and S W Karickoff, Sorption estimates for modeling, in Pesticides in the Soil Environment, Cheng, H H., Ed., Soil Science Society of America, Inc., Madison, WI, 79–101, 1990 46 Laird, D A et al., Adsorption of atrazine on smectites, Soil Sci Soc Amer J., 56 (1), 62–67, 1992 47 Wiedemeier T H et al., Natural Attenuation of Fuels and Chlorinated Solvents in the Subsurface, John Wiley & Sons, New York, 1999 48 McCarty, P L., M Reinhard, and B E Rittmann, Trace organics in groundwater, Environ Sci Techn., 15, 40–51, 1981 49 Dragun, J., The Soil Chemistry of Hazardous Materials, Hazardous Materials Control Research Institute, Silver Spring, MD, 1988 50 Hamaker, J W and J M Thompson, Adsorption in Organic Chemicals in the Soil Environment, Goring, C A I and J W Hamaker, Eds., Marcel Dekker, New York, 1972, 49–143 51 Herbert, B E et al., Pyrene sorption by water-soluble organic carbon, Environ Sci Technol., 27 (2), 398–403, 1993 52 Rao, P S C., L S Lee, and R Pinal, Consolvency and sorption of hydrophobic organic chemicals, Environ Sci Technol., 24 (5), 647–654, 1990 53 Chiou, C T and T D Shoup, Soil sorption of organic vapors and effects of humidity on sorption mechanism and capacity, Environ Sci Technol., 19, 1196–1200, 1985 ©2001 CRC Press LLC 54 MacIntyre, W G., T B Stauffer, and C P Antworth, A comparison of sorption coefficients determined by batch, column, and box methods on a low organic carbon acquifer material, Ground Water, 29 (6), 908–913, 1991 55 Tratnyek, P G and D L Macalady, Oxidation-reduction reactions in the aquatic environment, Handbook of Property Estimation Methods for Chemicals, Boethling, R S and D MacKay, Eds., Lewis Publishers, Boca Raton, FL, 2000 ©2001 CRC Press LLC ... congeners.19,34 ? ?20 01 CRC Press LLC CH3 OH 2H H OH O2 H TOLUENE CH2OH CH3 NO2 NO2 2- NITROTOLUENE CH2OH CH3 NO2 NO2 3-NITROTOLUENE CH3 CH3 CH3 OH OH + OH NO2 NO2 NO2 4-NITROTOLUENE O CCI2 CICH TCE Figure 2. 3a... anion (H2O2/HO2– ), hydroxyl radical (OH•), and ozone (O3).1 ,2, 55 O2 + e– + H+ = H2O˚ (protonated superoxide) (2. 29) O2 + 2e– + 2H+ = H2O2 (hydrogen peroxide) (2. 30) O2 + 3e– + 3H+ = H2O + HO˚... polychlorinated biphenyls (PCBs) by zero-valent iron with catalysis by Pd .2, 55 H CI C +H+ +2e-CI- C CI H CI C CI TCE Figure 2. 12 H H +H+ +2e-CI- C CI CI C cis-1 , 2- DCE +H+ +2e-CI- C H H H H C C H VC H Ethene