Environ. Sci. Technol. 2001, 35, 3163-3168
Impact of Peroxidase Addition on the Sorption-Desorption Behavior of Phenolic Contaminants in Surface Soils ALOK BHANDARI* AND FANGXIANG XU Department of Civil Engineering, Kansas State University, Manhattan, Kansas 66506-5000
The impact of peroxidase addition on sorption-desorption of phenol, o-cresol, 2,4-dichlorophenol (DCP), and 1-naphthol was evaluated using two surface soils. Target chemicals were added to soils as single solutes or binary mixtures. Seven-day adsorption studies were followed by sequential fill-and-draw extractions with synthetic groundwater. Addition of horseradish peroxidase (HRP) with H2O2 was the primary treatment evaluated. HRPmediated sorption enhancement was related to contaminant solubility and increased in the order: naphthol < DCP < cresol < phenol. Little or no competition was observed in the presence of cosolutes. Contaminant desorption from soils was dramatically reduced upon HRP addition. Reduction in desorption was quantified using the Hysteresis Index and interpreted as attenuation of contaminant mobility. Desorption data predicted that mobility reductions followed the order: naphthol < DCP < cresol < phenol. It is believed that enzyme addition resulted in the production of hydrophobic polymers that, due to their low aqueous solubilities, readily partitioned on to the solid-phase. The adsorbed polymers were less likely to partition into the aqueous phase than the parent phenols resulting in a reduced risk to the environment.
Introduction Soils contaminated with phenols pose a high risk to ecosystem health because of the multiple toxic effects associated with these chemicals at very low concentrations (1). Anthropogenic phenols enter the soil environment as a result of accidental spills and uncontrolled discharges; they may also accumulate as intermediates during the incomplete biodegradation of aromatic compounds and pesticide mixtures. Engineering remediation approaches to treat soils, sediments, and aquifer materials containing organic contaminants requires a thorough understanding of the governing physical, chemical, and biological processes in complex natural environments. Among these processes, the phenomenon of adsorption has often been defined as the most significant factor controlling the transport and ultimate fate of hydrophobic organic contaminants in natural and engineered systems (2-4). The adsorption-desorption behavior of organic contaminants in soils and sediments is greatly influenced by the amount and type of the associated natural organic matter (5-7). This soil/sediment organic matter (SOM) has been hypothesized to consist of amorphous and * Corresponding author phone: (785) 532 1578; fax (785) 532 7717; e-mail:
[email protected]. 10.1021/es002063n CCC: $20.00 Published on Web 07/04/2001
2001 American Chemical Society
condensed domains analogous to the rubbery and glassy phases of synthetic polymers that are capable of “absorbing” and “adsorbing” the organic solute, respectively (8). The equilibrium phase distribution of organic solutes between water and soil can be represented by the Freundlich isotherm model described by the following equations:
qe ) KFCne or log qe ) n log Ce + log KF where qe and Ce represent the solid-phase and aqueousphase concentrations of the target chemical, respectively. KF is a measure of the sorption capacity of the solid at a specific aqueous concentration of the solute. The exponent n is a joint measure of the magnitude and heterogeneity of energy associated with the sorption process. Values of n < 1, represent convex (type I) isotherms and are indicative of adsorption by heterogeneous media or condensed SOM where high energy sites are occupied first, followed by adsorption at sites with lower energies. Values of n > 1 represent concave (type III) isotherms that are indicative of adsorption behavior where previously adsorbed molecules modify the surface so that further sorption of the solute is enhanced. Finally, linear (type II) isotherms (n ) 1) represent a phenomenon whereby the organic solute partitions into an amorphous SOM structure in a manner analogous to the partitioning of the chemical from an aqueous phase into an organic solvent during liquid-liquid extraction. Several researchers have described the effect of SOM properties on the Freundlich isotherm coefficients, KF and n (4, 9, 10). In these studies, sorption of hydrophobic solutes to “young” organic matter, such as peat, was described by linear or near-linear isotherms, low sorption capacities (normalized to SOM content), rapid sorption, no competition from cosolutes, and low sorption hysteresis. Conversely, sorption to diagenetically “older” SOM, such as kerogen, demonstrated high sorption capacity and high sorption hysteresis, and was often nonlinear, slow, and competitive in the presence of cosolutes. Since organic matter associated with most soils falls in the geological time frame between “young” or peat-like, and “old” or kerogen-like, these geosorbents usually manifest sorption behaviors that are characterized in part by partitioning-type phase distribution and in part by site-specific adsorption. Huang et al. recently invoked the polymer sorption theory to propose the dual reactive domain model as an explanation for this type of sorption behavior in 27 different soils and sediments (11). Chemical or biochemical reactions between organic pollutants and soil components, such as those catalyzed by soil enzymes and transition metal oxides, can also exert significant influence on the ultimate fate of xenobiotics in soil systems (12-15). Soils contain a large background concentration of extracellular enzymes that mediate degradation and biosynthesis reactions in the soil environment. Several of these enzymes, including peroxidases, laccases, and polyphenol oxidases, can catalyze chemical reactions that result in the oxidative polymerization of phenolic chemicals (16-19). Such oxidative coupling reactions can result in the formation of covalent linkages between natural or anthropogenic phenolic chemicals and natural organic material, such as soil organic matter (SOM). Oxidative polymerization can also result in detoxification of substituted phenols as a result of fortuitous dehalogenation during the free-radical reaction of oxidative coupling (20). Past studies have demonstrated cross-coupling of substituted phenols, such as 2,4-dichlorophenol (DCP), to model humus constituents including orcinol, syringic acid, vanillic VOL. 35, NO. 15, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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acid, and vanillin (21). Laccase-catalyzed copolymerization of syringic acid was observed in the presence of a variety of halogenated phenols including 4-chlorophenol, 2,6-DCP, 4-bromo-2-chlorophenol, 2,3,6- and 2,4,5-trichlorophenol, 2,3,5,6-tetrachlorophenol, and pentachlorophenol (22). Oxidative coupling reactions catalyzed by three oxidoreductasess tyrosinase, peroxidase, and laccaseswere demonstrated for 2,4-DCP and a stream fulvic acid over a wide range of pH and temperature (23). In these studies, the polymerized product was observed to be less soluble than the parent phenol. Several recent studies investigating adsorption behavior of substituted phenols, anilines, hydroxylated polynuclear aromatic hydrocarbons (PAHs), and pesticides in soils and sediments have attributed the observed sorption hysteresis and irreversible binding of the contaminants to enzyme or transition metal oxide mediated oxidative coupling of the target chemicals to SOM (24-30). Nannipieri and Bollag (31) conducted a literature review and analysis on enzymemediated detoxification processes and presented a compelling argument for the use of enzymes for the clean up of pesticide-contaminated soils. Enzyme-mediated incorporation of phenolic chemicals into the SOM structure was proposed as a remediation approach capable of reducing the leachablility, toxicity, and bioavailability of the contaminant. This study was conducted to evaluate the feasibility of using the enzyme horseradish peroxidase to modify the behavior of phenolic chemicals and their mixtures in soils by altering their chemical structure. Mixtures were investigated because they are more representative of “real-world” contaminated sites, which rarely have single pollutants. It is argued that the resulting changes in contaminant behavior can have a significant impact on the ultimate fate of the phenolic contaminants in natural systems.
Experimental Section Soils. Two surface soils were collected near the city of Manhattan in Riley County, KS. These soils, obtained from an agricultural field (SOM ) 4.1%; pH ) 7.3) and an adjacent forested site (SOM ) 5.1%; pH ) 7.1), belong to the Haynie series and are classified as fine sandy loams. The site selection procedure for collecting soils was based on U.S. Department of Agriculture’s soil survey data and discussions with personnel at the Kansas Agricultural Experiment Station. Desired soil properties included near neutral pH, sandy/ silty composition, and varying organic matter content and type. The soils were collected aseptically and transported to the laboratory in coolers. Soils were sieved to pass through 1-mm and 500-µm sieves, and each soil was split into smaller representative fractions using the coning and quartering technique. The soil fractions were sterilized using a procedure that included subjecting the soils to sequential autoclaving and incubation to neutralize all spore-forming bacteria. Soil was autoclaved at 115 °C and 20 psi for 1 h and transferred into an incubator set at 37 °C for 2 days. The autoclaving and incubation procedure was repeated and followed by a final autoclaving to complete the sterilization procedure. When mixed with water, both soils were observed to release natural organic matter into solution. To avoid working with a three-phase system (soil, water, and dissolved organic material), the soils were washed several times with a synthetic groundwater solution (pH 7 phosphate buffer + 500 mg/L sodium azide, 18 mM ionic strength) to remove all leachable SOM. The washed soils were dried at 70 °C and homogenized with a mortar and pestle. The SOM content of the washed soils as determined by combustion at 550 °C was 2.5 and 3.7% for the agricultural and forest soils, respectively. The sand, silt, and clay fractions for the washed soils were 44, 42, 14%, and 56, 38, and 6% for the agricultural and forest soils, 3164
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TABLE 1. Selected Properties of the Target Chemicals
respectively. All prepared soils were stored at sub-zero temperatures in glass bottles. Chemicals. The target chemicals used in this study included phenol, o-cresol, 2,4-dichlorophenol, and 1-naphthol. Selected properties of these chemicals are listed in Table 1. Radiotracers were utilized to lower detection limits and track the distribution of the chemicals in the soil system. These tracers included 14C-phenol, 14C-o-cresol, 14C-2,4dichlorophenol (uniformly ring-labeled with specific activities of 14.3, 4.8, and 20.9 mCi/mmol, respectively), and 14C-1naphthol (carbon-1 labeled with specific activity of 9.6 mCi/ mmol). All target chemicals were purchased from Sigma Chemicals and used without further purification. All working solutions were prepared using the synthetic groundwater (GW) described previously. Nonlabeled target chemicals were added from concentrated stock solutions prepared in methanol. These solutions were then spiked with precise volumes of the corresponding radiotracer to obtain working solutions containing the desired amount of radioactivity. To avoid cosolvency effects due to methanol, the total volume of stock solutions added to GW was maintained at < 0.1% v/v. In studies involving mixtures, parallel experiments were conducted with one target chemical being radiolabeled while the other chemical was nonlabeled. Radioactivity in aqueous samples was enumerated as disintegrations per minute (dpm) using a Beckman 6500 liquid scintillation counter (LSC) with quench and luminescence corrections. Horseradish peroxidase (type II, RZ: 2.2) and hydrogen peroxide (30% w/w) were obtained from Sigma Chemicals and used without further purification. Adsorption. Preliminary investigations were conducted to determine the ideal solid/liquid ratio and solid/liquid contact time for adsorption experiments. Constant solid dose adsorption studies were conducted in glass centrifuge tubes operated as completely mixed batch reactors (CMBRs). The initial aqueous concentrations of the target chemicals were 5, 50, and 500 µM. These solutions were prepared by adding known amounts of stock solutions to synthetic groundwater using a Hamilton microliter syringe. Each CMBR contained 5 g (dry weight) of soil and approximately 14 mL solution. The experimental matrix included triplicate tubes consisting of (i) soil + chemical + enzyme + H2O2, designated as HRP; (ii) soil + chemical, designated as no HRP; and (iii) chemical only, designate as control. The “No HRP” treatment quantified sorption of the target chemical in the absence of enzyme. “Control” reactors were utilized to monitor contaminant losses due to volatilization or adsorption to reactor components. Preliminary experiments conducted with soil + chemical + H2O2 showed no changes in adsorption behavior of the target chemical, confirming the absence of Fenton’s oxidation in these systems.
TABLE 2. Freundlich Isotherm Parameters (n, KF, and R2) and Hysteresis Index (H. I.) for the Individual Chemicals Added to Agricultural and Forest Soils with and without Peroxidase Enzyme (HRP) chemical
HRP
phenol
FIGURE 1. Adsorption of target chemicals on agricultural soil: (a) individual chemicals: phenol (O), o-cresol (4), 2,4-DCP (0), and 1-naphthol (X); and (b) mixtures: phenol in phenol/naphthol mixture (O), phenol in phenol/DCP mixture (4), DCP in phenol/DCP mixture (0), and DCP in DCP/naphthol mixture (X). Each “HRP” tube was dosed with an excess of horseradish peroxidase, i.e., 2 activity units (AU)/mL of solution, with one AU being the amount of enzyme forming 1.0 mg of purpurogallin from pyrogallol in 20 s at pH 6.0 and 20 °C. Sufficient H2O2 was added to achieve a solution concentration equimolar to that of the target chemical(s). The tubes were capped with Teflon lined phenolic caps and the contents mixed with a vortex mixer. The tubes were placed in a endover-end tumbler and allowed to equilibrate at room temperature (22 ( 1 °C). A 7-day adsorption time was selected based on a preliminary rate study. At the end of the adsorption period, the tubes were removed from the tumbler, and their contents centrifuged at 250g for 45 min to separate the solid and aqueous phases. A 250-µL aliquot of the supernatant was removed from each CMBR and transferred into a minivial containing 5 mL of Fisher Scintisafe (50%) cocktail. The sample was then analyzed for radioactivity on the LSC. The disintegrations per minute (dpm) were translated to target chemical concentrations. Solid-phase concentration (qe, µmol/kg) was calculated from the difference between the initial aqueous concentration of the target chemical (C0, µM) and its concentration in water at the end of the equilibration period (Ce, µM). Data from the control tubes were used to adjust for abiotic losses. Desorption. Once the 250-µL aliquot was removed from the CMBRs at the end of the adsorption experiment, the soils were subjected to sequential “fill and draw” desorption using the procedure described in detail elsewhere (24). Briefly, this procedure consisted of removing the remaining supernatant from the test tube and replacing it with clean GW. The tubes were capped and placed in the tumbler for desorption. After 24 h, their contents were centrifuged, and the supernatant was drawn for analysis on the LSC. The aqueous phase from the test tube CMBRs was removed, the CMBRs were refilled with clean GW, and this “fill and draw” procedure was repeated until the radioactivity in the supernatant was reduced to below detection limit (50 dpm). Aqueous phase concentrations were calculated from the radioactivity measurements and solid-phase concentrations determined by performing a mass balance.
Results and Discussion Adsorption Behavior. The adsorption data obtained were fitted to the Freundlich isotherm model. Isotherms for phenol (O), o-cresol (4), 2,4-DCP (0), and 1-naphthol (X) sorption on agricultural soil are illustrated in Figure 1a. These isotherms represent adsorption behavior of individual solutes in the absence of HRP enzyme. The Freundlich sorption parameters n, KF, and R2 for the four chemicals and the two
no yes o-cresol no yes agricultural soil 2,4-DCP no yes 1-naphthol no yes phenol no yes o-cresol no yes forest soil 2,4-DCP no yes 1-naphthol no yes a
n (0.08)a
1.09 1.06 (0.07) 0.88 (0.09) 1.15 (0.09) 0.81 (0.03) 0.88 (0.05) 0.88 (0.12) 0.89 (0.08) 1.02 (0.10) 1.03 (0.06) 0.92 (0.08) 1.03 (0.09) 0.91 (0.03) 0.89 (0.04) 0.89 (0.05) 0.70 (0.03)
log KF
R2
H. I.
-0.84 0.13 (0.11) -0.3 (0.16) 0.33 (0.13) 0.66 (0.05) 0.72 (0.08) 1.26 (0.15) 1.36 (0.10) -0.65 (0.18) 0 (0.11) -0.24 (0.15) 0.21 (0.14) 0.77 (0.04) 0.65 (0.07) 1.07 (0.06) 0.88 (0.06)
0.994 0.990 0.988 0.993 0.999 0.996 0.976 0.990 0.989 0.996 0.990 0.991 0.999 0.997 0.997 0.997
0.53 4.86 0.28 2.12 0.30 1.23 0.66 0.06 0.96 2.18 0.01 1.65 0 1.24 0.86 0.86
(0.14)a
95% confidence interval.
soils are listed in Table 2. Isotherms were generally linear or near-linear with n for the four solutes ranging between 0.81 and 1.09 for the agricultural soil, and 0.89 and 1.02 for the forest soil. Although the forest soil had a higher SOM content, only minor differences were observed between n and log KF values for the two soils. Even though the SOM in most surface soils is predominantly “young”, the source of the soil, e.g., an agricultural field or a forest, can influence the ratio of the “young” to “old” organic matter types. In this study, sorption to forest soil appeared to be more linear for the phenols. As will be discussed later, naphthol sorption generally did not follow the trends observed for the phenols. The higher sorption linearity and lower sorption capacity (when normalized to SOM content) exhibited by the forest soil may be indicative of a larger component of the “younger” and more amorphous SOM in this soil. SOM in forest soils is usually rich in partially degraded organic material (lignin and cellulose) that is still in its initial stages of humification. Such organic matter is typically characterized by a high oxygen/ carbon (O/C) ratio and, a lower SOM hydrophobicity, translating to a reduced sorption capacity for organic solutes when normalized to SOM content (32). Younger SOM has been shown to behave like a rubbery polymer with its characteristic partitioning-type sorption behavior (8). Thus, when partitioning into the rubbery SOM domain is the dominant sorption process, it results in greater sorption linearity and n values that approach 1.0, as seen in the case of the forest soil (11). Figure 1b shows the effect of a cosolute on the adsorption behavior of phenol and DCP on the agricultural soil. The two lower isotherms in this figure represent phenol sorption in the presence of equimolar naphthol (O) or DCP (4), and the two upper isotherms are for DCP sorption in the presence of equimolar naphthol (X) or phenol (0). Isotherm parameters corresponding to these figures and those for the forest soil are listed in Table 3. The similarities between isotherms of phenol and DCP when present as single solutes (Figure 1a) or with cosolutes (Figure 1b) indicate no discernible competitive effects for the range of concentrations studied. Also, similar to the single solute systems, the sorption linearity in the two-solute systems was slightly higher in the forest soil as compared to the agricultural soil. Surface soils contain diagenetically “younger” organic matter rich in humic and fulvic acids, and organic matter derived from plant and animal material. Sorption of solutes to soils containing young and VOL. 35, NO. 15, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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TABLE 3. Freundlich Isotherm Parameters (n, KF, and R2) and Hysteresis Index (H. I.) for the Binary Mixtures Added to Agricultural and Forest Soils with and without Peroxidase Enzyme (HRP) mixture
phenol with no naphthol as cosolute yes phenol with no DCP as cosolute yes agricultural soil 2,4-DCP with no phenol as cosolute yes 2,4-DCP with no naphthol as cosolute yes phenol with no naphthol as cosolute yes phenol with no DCP as cosolute yes forest soil 2,4-DCP with no phenol as cosolute yes 2,4-DCP with no naphthol as cosolute yes a
N
HRP 0.91
(0.11)a
log KF -0.63
(0.19)a
R2
H. I.
0.985 0.50
1.01 (0.09) 0.04 (0.16) 0.989 2.10 0.88 (0.13) -0.56 (0.07) 0.993 0.74 0.98 (0.12) 0.74 (0.09)
0.28 (0.21) 0.980 1.71 0.56 (0.14) 0.989 0
0.91 (0.06) 0.79 (0.03)
0.66 (0.09) 0.995 1.04 0.52 (0.05) 0.998 0
1.11 (0.24) 0.49 (0.34) 0.950 0.82 0.93 (0.21) -0.60 (0.41) 0.940 0 1.09 (0.07) -0.37 (0.13) 0.964 2.15 0.91 (0.13) -0.53 (0.24) 0.975 0.05 0.92 (0.12) 0.87 (0.07)
0.09 (0.21) 0.980 1.85 0.56 (0.11) 0.999 0
0.82 (0.03) 0.89 (0.03)
0.76 (0.05) 0.998 0.38 0.65 (0.06) 0.998 0.06
0.88 (0.06)
0.67 (0.09) 0.995 1.04
95% Confidence Interval.
FIGURE 3. Effect of HRP addition on adsorption behavior target chemicals in agricultural soils: (a) individual solutes: phenol (O), o-cresol (4), 2,4-DCP (0), and 1-naphthol (X); and (b) mixtures: phenol in phenol/naphthol mixture (O), phenol in phenol/DCP mixture (4), DCP in phenol/DCP mixture (0), and DCP in DCP/naphthol mixture (X). hydrophobic solutes tend to associate preferentially with high-energy sites in soils resulting in site-specific adsorption, and a distribution of sorption energies (34). This distributed reactivity of the soil for more hydrophobic solutes is reflected by a decrease in the solute’s sorption linearity (n). Data plotted in Figure 2 also show that the sorption capacity factor increased directly with solute hydrophobicity and decreased logarithmically with solubility. Naphthol was consistently an outlier due to its lowest aqueous solubility, yet a smaller KOW than DCP. The trends illustrated in Figure 2 are consistent with those observed by other researchers (35). Effect of HRP on Adsorption. Figure 3 illustrates the effect of HRP addition on adsorption of the target chemicals when present as single solutes (Figure 3a) and when present as binary mixtures (Figure 3b). HRP addition resulted in significant enhancements in the sorption of all solutes. The extent of enhancement, however, was related to contaminant solubility and increased in the order: naphthol < DCP < cresol < phenol. It is believed that addition of HRP and H2O2 resulted in the production of oligomers as a result of oxidative polymerization of the target chemicals. The larger size and lower solubility of the polymers thus generated may have resulted in what appears to be enhanced adsorption of the target chemical. In reality, since we measured aqueous and solid phase radioactivity instead of the actual chemical, the sorption behavior shown in Figure 3 most likely relates to the adsorption of the polymer, not the parent target chemical, on the soil. Nevertheless, the net result of such a process is a reduction in contaminant mobility by elimination of the parent solute from the aqueous phase.
FIGURE 2. Impact of contaminant hydrophobicity and solubility on sorption of phenol, o-cresol, 1-naphthol, and 2,4-DCP on agricultural (-b-) and forest (- -O- -) soils: effect of log KOW on (a) Freundlich n and (b) log KF; and effect of contaminant solubility on (c) n and (d) log KF.
Sodium azide is a mechanism-based inactivator (suicide substrate) of horseradish peroxidase (36). It is likely that the presence of sodium azide at 500 mg/L may have interfered with the polymerization of phenols. The turnover of NaN3 by peroxidase can consume some of the available H2O2 and, therefore, result in partial polymerization of the available phenol. Results presented in this paper may, therefore, be conservative estimates of the extent of polymerization and subsequent immobilization achievable by peroxidase catalyzed oxidative coupling in systems without NaN3.
amorphous SOM (e.g., humic acid, fulvic acid, and peat) has been observed to be noncompetitive (33). Figure 2 describes the effect of solute properties (log KOW and aqueous solubility) on the sorption parameters, n and log KF. The curves in Figure 2 were drawn to illustrate general trends in the data. These trends show that the sorption behavior was controlled by contaminant properties. Sorption was generally more nonlinear for chemicals with higher KOW and lower solubility, naphthol being the exception. More
No discernible impact of the presence of cosolute on the effectiveness of HRP to enhance sorption of the target chemical was observed. Figure 3b illustrates the adsorption of phenol and DCP when present together as cosolutes, or when present with naphthol as a cosolute. Sorption parameters for the isotherms shown in Figure 3 are summarized in Tables 2 and 3. As seen from Figures 1 and 3, the general sorption behavior of DCP and naphthol appeared to remain unaltered upon HRP addition. The result was quite different
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for phenol and cresol where HRP addition resulted in significant enhancement in the sorption capacity (KF) and an increase in sorption linearity (n). This behavior can be attributed to the “salting out” or removal from solution of the large molecular weight polymers produced as a result of HRP mediated oxidative coupling of phenol and cresol. The polymers were able to readily “partition” from the aqueous phase on to the solid phase, and this behavior of the polymer was manifested by increases in the n and KF values. Although oxidative polymerization of DCP and naphthol also occurred (as evident by browning of the aqueous phase), the solubility difference between these target chemicals and their polymers may not have been as large as in the case of phenol and cresol. Enhancements in sorption linearity and capacity were highest for phenol and lowest for naphthol indicating the impact of the parent contaminant’s solubility on the effectiveness of HRP mediated polymerization. Desorption Behavior. The desorption behavior of the adsorbed solutes and their polymerization products was evaluated by conducting sequential “fill-and-draw” extractions with synthetic GW. Although data obtained from such experiments do not represent desorption equilibria, the results nevertheless describe the mass transfer characteristics of adsorbed solutes and provide important information about the relative ease with which the contaminants desorb from the solid phase. Desorption of organic contaminants from soils and sediments is often characterized by the occurrence of hysteresis, i.e., a desorption behavior different from adsorption. Possible reasons for sorption hysteresis include the presence of “ink-bottle” type pores that can trap the solute or, more likely, the occurrence of irreversible changes on the soil surface resulting in a desorption process that is actually different from adsorption (37). In this study, we quantified sorption hysteresis by a modified version of the Hysteresis Index (H.I.) parameter defined by Huang and Weber (38). These researchers characterized sorption reversibility by defining H.I. as:
H.I. )
FIGURE 4. Desorption of target chemicals from agricultural soil: (a) phenol (b, O) and cresol (2, 4), no HRP added; (b) DCP (b, O) and naphthol (2, 4), no HRP; (c) phenol (b, O) and cresol (2, 4) with HRP added; (d) DCP (b, O) and naphthol (2, 4) with HRP added. Solid symbols represent adsorption data and hollow symbols represent sequential desorption data.
qde - qae |T, Ce qae
where qae and qde are solid-phase solute concentrations for the adsorption and desorption experiments, respectively, and T and Ce specify conditions of constant temperature and residual aqueous phase concentration. In this study, we used the first set of our sequential desorption data points to develop a desorption “isotherm” (not shown). This was clearly not an isotherm in the real sense because the 24 h contact time allowed to acquire the first set of desorption data points was not believed to represent equilibrium conditions. However, this approach was considered satisfactory to compare the ease of desorption of various solutes in the presence and absence of HRP. In our approach, a H. I. ) 0 was indicative of no sorption hysteresis. H. I. values were calculated at Ce ) 10 µM. Data obtained from the sequential desorption experiments on agricultural soil is presented in Figures4 and 5. The corresponding adsorption isotherms are also shown. Figure 4a,b illustrate desorption behavior of the individual target chemicals in the absence of HRP, and Figure 4c,d illustrates corresponding desorption behavior when HRP was added to the CMBRs. Similarly, Figure 5a,b show desorption of phenol with naphthol and DCP as cosolutes, and DCP with naphthol and phenol as cosolutes, in the absence of HRP. Figure 5c,d illustrates the impact of HRP addition on desorption of phenol and DCP when present with cosolutes. The H. I. values for data presented in Figures 4 and 5, and for the corresponding data for the forest soil are tabulated in Tables 2 and 3.
FIGURE 5. Desorption behavior of mixtures in agricultural soil: (a) phenol with naphthol (b, O) and phenol with 2,4-DCP (2, 4), no HRP added; (b) DCP with phenol (b, O) and DCP with naphthol (2, 4), no HRP; (c) phenol with naphthol (b, O) and phenol with 2,4DCP (2, 4) and HRP added; (d) DCP with phenol (b, O) and DCP with naphthol (2, 4) and HRP added. Solid symbols represent adsorption data and hollow symbols represent sequential desorption data. H. I. values illustrate negligible sorption hysteresis for individual target chemicals in the absence of HRP. Calculated H. I. values were 0.53, 0.28, 0.30, and 0.66 for phenol, cresol, DCP, and naphthol, respectively. Figure 4 confirms these VOL. 35, NO. 15, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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numbers, except in the case of naphthol where a large sorption hysteresis was observed despite its small calculated H. I. value. The impact of HRP addition on desorption behavior of the target chemicals is apparent in Figure 4c,d. Desorption of the three phenols was dramatically reduced in the presence of HRP. H. I. values for phenol, cresol, and DCP in the presence of HRP were 4.86, 2.12, and 1.23, respectively. The H. I. values were correlated to the aqueous solubilities of the target chemicals. More effective contaminant immobilization occurred for chemicals with higher mobility. It is believed that the lower solubility of the polymers generated during HRP catalyzed polymerization of the phenols resulted in the reduced desorption. The desorption behavior shown in Figure 4 most likely relates to desorption of the polymer and not the parent target chemical. Nevertheless, HRP addition appeared to result in the immobilization of the parent phenol by transforming the contaminant into hydrophobic macromolecules that adsorbed strongly to the solid phase. Once again, naphthol behaved differently from the three phenols. No increase in hysteresis was observed upon HRP addition. Desorption of phenol and DCP in the presence of cosolutes is illustrated in Figure 5. Cosolutes had little or no effect on desorption behavior of phenol and DCP when no HRP was added. Little or no hysteresis was observed for phenol and DCP in the presence cosolutes. Hysteresis was dramatically enhanced upon HRP addition with H. I. values for phenol with naphthol and DCP as cosolutes calculated to be 2.10 and 1.71, respectively. Addition of HRP also increased the H. I. for DCP with phenol as cosolute from 0 to 1.04, and with naphthol as cosolute from 0 to 0.82. It was also observed that the H. I. values in HRP amended systems were lower with cosolutes than when the target chemicals were present as the sole solutes. This appears to indicate that the presence of co-contaminants in the soil may adversely impact peroxidase-catalyzed immobilization of phenols in soils. Nevertheless, HRP addition did result in a significant reduction in contaminant leachability even in the presence of cosolutes, indicating the effectiveness of this immobilization process for surface soils contaminated with phenolic mixtures. Results from this study demonstrate the effectiveness of peroxidase in reducing the mobility of phenolic contaminants in surface soils. This reduction in mobility is brought about by the polymerization of phenols and subsequent adsorption of the polymers on soil particles. The large size and low aqueous solubilities of the polymers prevent these macromolecules to desorb readily. It is also likely that enzymemediated oxidative polymerization may have resulted in the direct cross-linking of the target compounds to SOM. Such reactions can attenuate the migration of phenolic pollutants in soils by reducing contaminant transport with surface runoff or leaching into the subsurface. The more soluble and, therefore, more mobile phenols are affected to a greater extent by the peroxidase mediated immobilization technique. The technique remains highly effective even in the presence of other phenols or hydroxylated PAHs. The behavior of hydroxylated PAHs such as naphthol, however, cannot be accurately predicted from trends observed for phenols.
Acknowledgments The authors acknowledge financial support of the National Science Foundation (Award Numbers EPS-9874732 and BES9875241), the Kansas Agricultural Experiment Station (Con-
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tribution Number 01-270-J), and the Kansas Technology Enterprise Corporation. Inkwon Cho, Prachi Mungali, and Ryan Harvey assisted with the experimental work. Heather Lesan provided detailed comments on the manuscript.
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Received for review December 29, 2000. Revised manuscript received April 9, 2001. Accepted May 21, 2001. ES002063N
