Competition for Sorption and Degradation of Chlorinated Ethenes in

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Environ. Sci. Technol. 2004, 38, 2879-2884

Competition for Sorption and Degradation of Chlorinated Ethenes in Batch Zero-Valent Iron Systems J A N D R I E S , †,‡ L E E N B A S T I A E N S , * ,† D I R K S P R I N G A E L , †,§ SPIROS N. AGATHOS,‡ AND LUDO DIELS† Department of Environmental Technology, Flemish Institute for Technological Research, Boeretang 200, 2400 Mol, Belgium, and Unit of Bio-engineering, Catholic University of Louvain, Louvain-La-Neuve, Belgium

The sorption and degradation of the chlorinated ethenes tetrachloroethene (PCE, 5 mg L-1) and trichloroethene (TCE, 10 mg L-1) were investigated in zero-valent iron systems (ZVI, 100 g L-1) in the presence of compounds common to contaminated groundwater with varying physicochemical properties. The potential competitors were chlorinated ethenes, monocyclic aromatic hydrocarbons, and humic acids. The effect of a complex matrix was tested with landfill contaminated groundwater. Nonlinear Freundlich isotherms adequately described chloroethene sorption to ZVI. In the presence of the more hydrophobic PCE (5 mg L-1), TCE sorption and degradation decreased by 33% and 30%, respectively, while TCE (10 mg L-1) decreased PCE degradation by 30%. In the presence of nonreactive hydrophobic hydrocarbons (i.e., benzene, toluene, and m-xylene at 100 mg L-1), TCE and PCE sorption decreased by 73% and 55%, respectively. The presence of the hydrocarbons had no effect on TCE degradation and increased PCE reduction rates by 50%, suggesting that the displacement of the chloroethenes from the sorption sites by the aromatic hydrocarbons enhanced the degradation rates. Humic acids did not interfere significantly with chloroethene sorption or with TCE degradation but lowered PCE degradation kinetics by 36% when present at high concentrations (100 mg L-1). The landfill groundwater with an organic carbon content of 109 mg L-1 C had no effect on chloroethene sorption but inhibited TCE and PCE degradation by 60% and 70%, respectively.

Introduction Numerous research groups have carried out investigations to improve our understanding of the iron/contaminant interactions in zero-valent iron (ZVI) systems. When chlorinated aliphatics in solution come in contact with ZVI, they undergo a thermodynamically favorable reductive dechlorination that is equivalent to iron corrosion with the chlorinated aliphatic serving as the oxidizing agent. The three major reductants in ZVI systems are (i) the iron metal Fe(0), * Corresponding author e-mail: [email protected]; telephone: +32 14-33-51-79; fax: +32 14-58-05-23. † Flemish Institute for Technological Research. ‡ Catholic University of Louvain. § Present address: Laboratory for Soil and Water Management, Catholic University of Leuven, Leuven, Belgium. 10.1021/es034933h CCC: $27.50 Published on Web 04/17/2004

 2004 American Chemical Society

(ii) ferrous iron associated with iron oxide coatings, and (iii) hydrogen in the presence of an appropriate catalyst (1-4). Previous research supports reductive dechlorination directly coupled to the oxidative dissolution of the zerovalent metal as the dominant mechanism (3, 5). The electrons are transferred from the ZVI surface to the chlorinated aliphatic. The surface of ZVI in water is coated with a film of mixed-valent iron oxides, the most prevailing being magnetite (5, 6). The electron transfer can occur either through the conductive magnetite film, in the numerous pores of the magnetite coating, or in the “defects” or “abnormalities” present at the surface (3, 6-8). Iron dissolution at the bottom of the pores or defects and chloroethene reduction at the mouth would be the controlling anodic and cathodic processes, respectively. Another possibility is reduction of chlorinated aliphatics at the surface of the iron oxide coating by electrons transferred from surface-bound ferrous iron (5). From the considerations above, it is clear that the reductive dechlorination of chlorinated compounds by ZVI is a heterogeneous or surface-mediated reaction. Such reactions require that the reactants reach the solid surface, where they can interact with reactive or nonreactive sites (9). Chemical reactions (i.e., breaking of bonds) occur at the reactive sites, and nondestructive sorption occurs at the nonreactive sites. Burris et al. (10) investigated sorption of trichloroethene (TCE) and tetrachloroethene (PCE) to ZVI in closed batch systems by determining aqueous and sorbed concentrations as a function of time. Both compounds exhibited substantial nonlinear sorption. The sorption behavior of chloroethenes to different types of cast iron could be described by a generalized Langmuir isotherm expression but also by a nonlinear Freundlich equation (10-12). The nonlinearity of chlorinated ethenes sorption to cast iron implies the existence of either a finite number of sorption sites or sites with different sorption energies. The order of sorption capacity was PCE > TCE > trans-dichloroethene (transDCE) > cisDCE. The majority of the sorption of the chlorinated ethenes was to the embedded graphite nodules or flakes. The graphite carbon represents the nonreactive sorption sites on ZVI (10-12), although a recent study demonstrated that exposed graphite also mediated reduction of 2,4-dinitrotoluene (13). It still remains to be determined whether chloroethene degradation can be mediated by graphite. Many studies report that the degradation of chlorinated organic compounds by ZVI proceeds by kinetics that are first-order with respect to contaminant concentration (14, 15). There is however growing evidence that observed firstorder rate “constants” sometimes significantly decrease with increasing concentrations of the chlorinated organic (2, 16). Such behavior is indicative of surface reaction limitation or reactive site saturation (11). The existence of a finite number of reactive and sorption sites implies that competition for reaction and sorption, respectively, may occur under certain conditions (2, 7, 10, 16-18). Groundwater pollution (e.g., in a landfill leachate contaminated groundwater) may consist of complex mixtures of compounds (19). Groundwater co-contaminants with different physicochemical properties may impact the efficiency of a ZVI treatment by interacting with processes at either or at both the reaction and the sorption sites on the iron surface. To test this hypothesis, we investigated the sorption and degradation of the chlorinated ethenes PCE and TCE in ZVI systems in the presence of different classes of compounds. The compounds selected were expected to compete with the chlorinated ethenes at different sites on VOL. 38, NO. 10, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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the iron surface, depending on their physicochemical characteristics. The potential competitors examined were reactive chlorinated ethenes, nonreactive moderately hydrophobic monocyclic aromatic hydrocarbons, and hydrophilic humic acids (HAs). The effect of a complex matrix, a landfill leachate contaminated groundwater, on chlorinated ethene sorption and degradation was tested.

Experimental Section Media. Unless otherwise stated, all experiments were performed with a dilute simulated groundwater at neutral pH. The simulated groundwater consisted of NaHCO3, KHCO3, CaCl2‚2H2O, and MgCl2‚6H2O at a concentration of 0.5 mM each in Milli-Q water. In experiments with HAs, commercially available sodium salts of HAs were used, supplied by Aldrich. The HAs were used without pretreatment. The ZVI was cast iron supplied by Gotthart Maier Metallpulver (Germany). The iron filings had a size ranging from 0.25 to 2 mm and a specific surface area of 0.745 ( 0.007 m2 g-1 (as determined by N2 single-point BET analysis). The total carbon content was 2.8%, consisting of embedded graphite flakes, typical of cast iron (12). The chemicals PCE (99.9% purity), TCE (>99.5%), benzene (>99.7%), and toluene (>99.8%) were supplied by Merck (Germany), and m-xylene (> 99%) was purchased from Janssen Chimica (Belgium). Setup of the Batch Experiments. Five series of batch experiments were set up. In a first test, the sorption and degradation of PCE (at 5 mg L-1) and TCE (at 10 mg L-1) and the sorption of benzene, toluene, and m-xylene were examined. In the second test, the competition between PCE and TCE for sorption and degradation was evaluated. In subsequent tests, we investigated the effect of the following constituents on the degradation and sorption of PCE (at 5 mg L-1) and TCE (at 10 mg L-1): (i) mixtures of the aromatic hydrocarbons benzene, toluene, and m-xylene (BTX) at low (10 mg L-1 each) and high concentrations (100 mg L-1 each); (ii) Aldrich HAs at low (20 mg L-1) and high concentration (100 mg L-1); and (iii) a landfill leachate contaminated groundwater. The experiments testing competition effects with chlorinated ethenes and aromatic hydrocarbons were performed twice to investigate the quality of the reported data. Each test consisted of a series of 26-mL glass serum flasks supplied with 100 g L-1 of ZVI (corresponding to a surface area of 74.5 m2 L-1). The flasks were filled, leaving no headspace, with simulated groundwater (except in tests where the landfill groundwater sample was used) containing the compound(s) of interest. TCE (at 10 mg L-1) and PCE (at 5 mg L-1) were supplied from water-saturated stock solutions. The BTX compounds were supplied as pure phase chemicals. Simulated groundwater solutions containing HAs were filtered over folded filter paper (Whatman) before the addition of TCE or PCE. The reaction vials were crimp-sealed with Teflon-faced rubber septa and incubated at room temperature on a rotary shaker (10 rpm). Selected reaction vials were removed and sacrificed for analysis after 1, 2, 3, 4, 7, and 10 d. Samples were taken for determination of pH, oxidation-reduction potential (ORP), and the concentration of aqueous and sorbed chlorinated ethenes and, when applicable, aromatic hydrocarbons and humic acids. Samples for the determination of HAs were filtered (0.45 µm) before analysis. The landfill groundwater sample was collected down gradient of a municipal landfill near Antwerp (Belgium) and was characterized by a high specific conductivity (3150 µS cm-1), a relatively high concentration of dissolved organic carbon (DOC: 109 mg of C L-1), and a neutral pH. The groundwater sample was filtered over a membrane filter (0.45 µm) before use. Zero-Valent Iron Extraction Protocol. To determine the sorption isotherms for TCE and PCE and to estimate the 2880

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sorption kinetics of the individual BTX compounds in ZVI systems, we determined the evolution of both their aqueous and sorbed concentrations (10, 11). The following protocol, modified from Allen-King et al. (11), was applied. First, a magnet was used to clear the solution from suspended ZVI filings. Samples were then taken from the aqueous phase for pH, ORP, BTX, and chlorinated ethene analyses, yielding the aqueous concentration C. The caps and septa were removed, and as much as possible of the remaining aqueous phase was removed with a syringe. By weighing the serum bottles before and after removal of the liquid phase, we found that approximately 98 ( 1% of the solution was removed. The empty serum vials containing the ZVI filings were immediately filled with pure methanol, leaving no headspace, crimp-sealed with Teflon-faced rubber septa, and equilibrated on a rotary mixer (10 rpm) to extract the sorbed contaminants. After 30 min, the iron was allowed to settle (using a magnet), and the supernatant was sampled, yielding the sorbed concentration q. The methanol samples were diluted 10-fold to prevent interference with the GC/MS headspace analysis. Using this procedure, a high recovery was achieved for the conservative (i.e., nonreactive) compounds benzene (100 ( 2% recovery, based on triplicate setups), toluene (98 ( 3% recovery), and m-xylene (94 ( 7% recovery). To determine the sorption isotherms for the individual BTX compounds in ZVI systems, a series of duplicate iron-containing batches was set up supplied with the BTX compounds in a concentration range spanning 3 orders of magnitude (from 10 µg L-1 to 10 mg L-1). After 3 d of equilibration, the aqueous concentration was determined using the procedure described above. The sorbed concentration was estimated by subtracting the aqueous concentration measured in ZVI-containing batches from the aqueous concentration determined in blanks without ZVI. Analyses. The pH was measured with a Hamilton glass combination electrode and a WTW pH 325 pH meter. The ORP was measured with a WTW Sentix glass combination electrode connected to a Russel meter type RL250. The specific conductivity was measured with a WTW Tetracon 325 electrode connected to a WTW multiline P4 universal meter. The concentration of chlorinated ethenes and aromatic hydrocarbons was determined with an Interscience GC 800 top gas chromatograph connected to a Voyager MS detection system. The samples were taken from the headspace of the sample vials after equilibration for 30 min at 50 °C using a headspace autosampler (CombiPal CTC Analytics). The column was a DB5MS (60 m length, 0.25 mm i.d., and 0.25 µm film thickness). The flow rate of the carrier gas (helium) was 0.9 mL min-1. The temperature program started at 40 °C for 3 min, ramped at 5 °C min-1 up to 150 °C, and at 25 °C min-1 to 250 °C. Calibration was performed in the 0-5 mg L-1 range using a mixture of toluene-d8 and ethylbenzene-d10 as internal standard. The concentration of humic acids was determined with a UV/visible spectrophotometer Ultro Spec 3000 at a wavelength of 254 nm. Modeling. A pseudo-first-order model (eq 1) was applied to describe the reductive dechlorination of a parent compound by ZVI (14):

C ) C0e-kt

(1)

with C as the concentration at any time, C0 as the initial concentration, k as a first-order rate constant (h-1), and t as the reaction time. The natural logarithmic transformation of eq 1 yields a linear equation with the first-order rate constant k as slope. The k values were estimated by linear regression of the transformed data versus time using MS Excel’s data analysis tools (Microsoft Corporation, Redmond, WA). The first-order rate constant was normalized to a total iron surface

FIGURE 1. Sorption isotherms of PCE (9) and TCE (0) in zero-valent iron batch systems, normalized on the basis of aqueous solubility (q, sorbed concentration; C, aqueous concentration). The fitted Freundlich sorption isotherm coefficients are found in Table 1. area of 1 m2 mL-1 to facilitate comparison with the kinetic parameters reported in other studies. The equilibrium distribution of organic compounds between aqueous and solid phases in ZVI systems was described by a Freundlich sorption isotherm expression (eq 2) (20):

q ) K FC N

(2)

with q and C as the sorbed (µmol kg-1) and aqueous (µM) equilibrium concentration, respectively; KF as the Freundlich sorption capacity ((µmol kg-1)/(µM)N), and N as the degree of isotherm linearity. The Freundlich parameters KF and N were estimated by linear regression analysis of the natural log-transformed experimental data. The area below the sorption isotherm was computed by integration of eq 2.

Results and Discussion Sorption of TCE and PCE to Zero-Valent Iron. The setup of each of the series of batch experiments included reference sets where the reactive flasks (i.e., with ZVI) were supplied with only PCE or TCE, without amendments. The partitioning of the aqueous and sorbed TCE and PCE concentrations for each of these sets is shown in Figure 1. The aqueous phase concentrations were normalized to the aqueous solubilities of the respective compounds (25 °C; PCE: 0.904 mM; TCE: 8.365 mM) (21). The sorption isotherm for PCE and TCE overlap after normalization to the aqueous solubility, in accordance to Traubes rule for sorption of hydrophobic compounds to carbon, suggesting that the graphite present in Gotthart Maier cast iron was the main sorbent (12). Table 1 shows the estimated Freundlich sorption isotherm coefficients for PCE and TCE in the reference sets. The data could be described quite well by nonlinear Freundlich sorption isotherm expressions (Table 1: R 2 > 0.96). Nonlinearity, which is indicated by a coefficient N different from 1, is typically observed for sorption of hydrophobic organic compounds to so-called “hard carbon” such as graphite (12, 22). The sorption data should be considered as quasi-equilibrium sorption isotherms since chemical degradation reactions are also taking place (10, 12). The Freundlich sorption capacity KF for PCE sorption was 4 times higher than the KF value for TCE sorption, indicating the higher affinity of ZVI for PCE sorption. This was also observed by Burris et al. (12) for sorption of chloroethenes to four different types of cast irons. The parameters estimated in our experiments are below the range of Freundlich sorption isotherm coefficients reported by Burris et al. (12) (Table 1), which may indicate that the cast iron in the present study is a less efficient sorbent. Burris and co-workers (12) found that the various sorption data did not correlate well with the carbon content and surface area of the cast irons, suggesting that other ZVI-specific characteristics, such as the exposed graphite surface area, may control the extent of sorption.

Sorption of BTX Compounds and Humic Acids to ZeroValent Iron. The aqueous BTX concentrations in ZVI systems rapidly declined to reach equilibrium after approximately 24 to 48 h (results not shown). Increasing the amount of ZVI led to lower equilibrium concentrations. Sorption equilibrium was adequately described by nonlinear Freundlich isotherms (Table 1: R 2 > 0.97). The extent of sorption was highest for m-xylene and lowest for benzene. The series of KF values for the chlorinated ethenes and aromatic hydrocarbons correlate quite well with their respective octanol-water coefficients (R 2 ) 0.899), indicating the hydrophobic nature of the sorption process in ZVI systems. Due to the hydrophobicity of the BTX compounds, it is very likely that most BTX sorption occurred on the embedded graphite carbon in the cast iron, similar to PCE and TCE. In the experiments with HAs, we observed that the UV absorbance at 254 nm, which is a measure for the HAs concentration (23), remained constant in the control sets without ZVI but rapidly decreased to background levels in the reactive sets with ZVI (results not shown). The HAs were thus completely removed from the aqueous phase. Humic acids are soluble polyelectrolytes, and their mechanism of sorption is probably completely different from the hydrophobic interactions occurring with BTX and chlorinated ethenes. Sorption of negatively charged ligands such as HAs is assumed to occur through a ligand exchange mechanism with the hydroxylated surface groups on the iron oxide coating (20, 24). Competition for Sorption. Table 2 summarizes the Freundlich isotherm coefficients for PCE and TCE sorption in the presence of chlorinated ethenes, mixed aromatic hydrocarbons, and HAs. The estimated KF and N values varied considerably. The area below the sorption isotherm, computed by integration of the fitted Freundlich expression, was therefore used as a measure for the magnitude of sorption for each amendment. Figure 2 illustrates the effect of the different amendments on the sorption of TCE to ZVI. Changes of more than 25% and 50% were considered as moderate and significant, respectively. The following patterns emerge from Figure 2 and Table 2: (i) TCE had no effect on PCE sorption, but PCE lowered TCE sorption by 33%; (ii) the presence of mixed BTX significantly lowered TCE sorption (by about 73%) and to a lesser extent PCE sorption (up to 55%); (iii) humic acids only slightly reduced TCE and PCE sorption (by less than 22%). The landfill groundwater did not affect chloroethene sorption (results not shown). In a study by Burris et al. (10), PCE affected TCE sorption and vice versa. The difference between our results and the observations from Burris and co-workers (10) may be due to the higher contaminant loadings used in their study (about 2-fold). The fact that PCE appears to affect sorption of TCE more than vice versa, as observed in our experiments (Table 2), is in agreement with the higher sorption affinity of graphite for the more hydrophobic PCE (Table 1) (22). Differential hydrophobicities may also explain the differences in the relative magnitude of displacement of PCE and TCE by the BTX compounds (Table 2). Competition for hydrophobic sorption sites is expected to mostly affect the compound with a lower affinity (i.e., TCE) (22). Higher concentrations of the more hydrophobic aromatic hydrocarbons toluene and m-xylene probably effectively competed with the less hydrophobic TCE. Finally, the fact that binding of humic acids to the ZVI surface did not significantly affect chloroethene sorption can be explained by the difference in sorption sites for the two compounds as discussed above. The small impact observed (4-22% lower sorption) might result from the fact that some sorption of chloroethenes did not happen to the graphite inclusions but at the iron oxide surface where humic acids are expected to interact (23, 24). In one of the few studies in this field, Tratnyek et al. (25) observed that VOL. 38, NO. 10, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 1. Estimated Freundlich Sorption Isotherm Coefficients for Sorption of Chlorinated Ethenes PCE and TCE and Aromatic Hydrocarbons to Zero-Valent Cast Iron (100 g L-1) with a Carbon Content of 2.8%a compound

log Kow b (25 °C)

reaction time (d)

R 2 (n)

KF (µmol/kg)/ (µM)N

N

L-1)c

2.88

1-10

2.42

1-10

2.13 2.69 3.15

3 3 3

0.96 (17) 0.801-0.96d 0.97 (13) 0.858-0.979d 0.97 (12) 0.98 (14) 0.99 (12)

32 (1) 72.9-164d 8 (1) 12.3-18.7d 7 (1) 18 (1) 46 (1)

0.52 (0.03) 0.21-0.45d 0.70 (0.03) 0.36-0.55d 0.63 (0.03) 0.69 (0.03) 0.63 (0.02)

PCE (5 mg

TCE (10 mg

L-1)c

benzene (0.01-10 mg L-1)c toluene (0.01-10 mg L-1)c xylene (0.01-10 mg L-1)c

a n, number of data points; K , Freundlich sorption capacity; N, Freundlich degree of isotherm linearity. The standard error is shown in parentheses. F Octanol-water partition coefficient reported by Schwarzenbach et al. (21). c Initial concentration or concentration range. d Range of Freundlich sorption isotherm coefficients reported by Burris et al. (12) for four different cast irons with a carbon content of 2.8-3.2%. b

TABLE 2. Confidence Intervals (95%) for Estimated Freundlich Sorption Isotherm Coefficients for Sorption of Chlorinated Ethenes PCE and TCE to Zero-Valent Cast Iron and Relative Change of Sorption Magnitude in the Presence of Chlorinated Ethenes, Aromatic Hydrocarbons, or Humic Acidsa PCE (5 mg L-1)

TCE (10 mg L-1)

addition

KF

N

A/Aref

no addition chlorinated ethenesc 10 mg L-1 BTX 100 mg L-1 BTX 20 mg L-1 HA 100 mg L-1 HA

29-35 27-32 15-22 3-6 21-35 37-63

0.46-0.57 0.46-0.58 0.54-0.86 0.86-1.36 0.35-0.59 0.15-0.42

1 0.93 0.82 0.45 0.78 0.96

b

KF

N

7-9 5-6 1-2 0.7-1.4 8-13 7-12

0.62-0.78 0.63-0.76 0.79-1.01 0.85-1.09 0.45-0.66 0.47-0.71

A/Aref b 1 0.67 0.31 0.27 0.87 0.85

a K , Freundlich sorption capacity; N, Freundlich degree of isotherm linearity. b Comparison of the area below the isotherm for a certain addition F (A) and for the reference case without addition (Aref), calculated by integration of the fitted Freundlich expression within the experimental range (0-15 µM for PCE and 0-30 µM for TCE). c TCE (at 10 mg L-1) or PCE (at 5 mg L-1).

FIGURE 2. Quasi-equilibrium sorption isotherms of TCE (at 10 mg L-1) in the presence of PCE, BTX, or humic acids in batch zerovalent iron systems: (×) PCE (at 5 mg L-1); (9) BTX mixture at 10 mg L-1 each; (0) BTX mixture at 100 mg L-1 each; (2) humic acids at 20 mg L-1; (4) humic acids at 100 mg L-1; (s) Freundlich model fitting for the reference system without addition (q, sorbed concentration; C, aqueous concentration). The Freundlich sorption isotherm coefficients are found in Table 2. Suwannee River natural organic matter (at 40 mg L-1) decreased TCE sorption most probably as a result of competitive adsorption. Tratnyek et al. (25) determined TCE adsorption isotherms over a much wider concentration range (up to 6 mM) and at a TCE loading approximately 20 times higher than in our study. Under these circumstances, it can be expected that a larger fraction of TCE would also sorb to the iron oxide coating where competitive adsorption with humic acids would occur. Differences in the type of HAs in the two studies may also play a role in these discrepancies. Two additional but distinct processes (i.e., enhanced solubilization or enhanced sorption) can result from the retention of hydrophobic compounds within hydrophobic moieties of the dissolved or sorbed humic acids, respectively (21, 26, 27). Since the humic acids were completely removed from the aqueous phase in our experiments, the solubilization effect could not play a role. Enhanced chloroethene sorption also was not observed (Table 2). Competition for Reaction. The surface area normalized first-order rate constant, estimated from the lumped reference degradation data (i.e., without additions) for the different 2882

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experimental sets, was 0.11 ( 0.01 h-1 for PCE (n ) 19; R 2 ) 0.76) and 0.21 ( 0.02 h-1 for TCE (n ) 17; R 2 ) 0.86). These kinetic parameters were based on total concentrations (i.e., aqueous + sorbed). When based on aqueous concentrations only, the surface area normalized first-order rate constants were 0.20 ( 0.03 h-1 for PCE (86% higher) and 0.25 ( 0.03 h-1 for TCE (20% higher). The overestimation of kinetic parameters based on aqueous concentrations only was thus more pronounced for PCE than for TCE, as observed previously (10, 12). When taking sorption losses into account, the degradation rate of PCE was about half that of TCE. Similar results were observed by others (9, 16). This is probably due to the fact that a greater portion of PCE is sorbed to nonreactive sites and does not take part in degradation reactions. Concurrent with chloroethene degradation, a typical increase in pH (above 9.0) and a decrease in ORP (below -300 mV) were observed (results not shown). Figure 3 shows the effect of chlorinated ethenes, aromatic hydrocarbons, humic acids, and a landfill groundwater on the degradation kinetics of PCE and TCE by ZVI. The displayed kinetic coefficients were based on total concentrations and expressed relative to the first-order rate constant of the reference without amendments within that specific experimental set. The degradation rates of both PCE and TCE decreased by 30% in dual-component systems (i.e., TCE and PCE supplied together) as opposed to single-component systems (Figure 3). The presence of the mixed BTX compounds at low (10 mg L-1) and high (100 mg L-1) concentrations increased the PCE degradation rate by approximately 50%. The presence of the humic acids at low (20 mg L-1) and high (100 mg L-1) concentrations decreased the PCE degradation rate by approximately 13% and 36%, respectively. The effect of any of these additions on TCE removal was not significant and less than 15% (Figure 3). The most pronounced effect was observed in the presence of the landfill groundwater: PCE and TCE degradation kinetics decreased by 70% and 60%, respectively.

FIGURE 3. Effect of liquid amendments and composition on the first-order degradation rate constant k relative to the reference rate constant k - ref (without amendments) for TCE and PCE degradation in zero-valent iron batch systems: effect of PCE (5 mg L-1) or TCE (10 mg L-1), of low (10 mg L-1 each) and high (100 mg L-1 each) BTX concentrations, of low (20 mg L-1 HA) and high (100 mg L-1 HA) humic acids concentrations, and of a landfill leachate contaminated groundwater (error bars represent one standard deviation). We observed significant inter-species competition between PCE and TCE for reaction (Figure 3), implying a finite number of reaction sites. Both PCE and TCE were affected by the competition for reaction sites, while only TCE was affected by competition for sorption sites (Figure 2), suggesting that reaction and sorption sites are different physicochemical entities on the ZVI surface. Burris et al. (10) observed competition between TCE and PCE for sorption but not for reaction. On the contrary, Arnold and Roberts (16) found substantial competition and saturation effects with chlorinated ethenes in batch ZVI systems. In column systems, saturation effects with TCE were sometimes observed but not always (7, 17, 18, 28). Interspecies competition between cisDCE and TCE was substantial in column systems: TCE significantly lowered DCE degradation rates, while TCE removal was not substantially affected by DCE (17, 18). To our knowledge, this is the first report on the effects of nonreactive aromatic hydrocarbons (BTX) on the ZVI process. Mixed BTX decreased PCE sorption (Table 2) and significantly enhanced PCE degradation (Figure 3). The explanation we propose is that dislocation of sorbed chloroethenes from nonreactive sorption sites by hydrophobic co-contaminants such as BTX increases the degradation kinetics of the chlorinated ethenes. Earlier studies suggested that a decrease of the chloroethene concentration at the ZVI surface, due to the presence of surfactants and cosolvents, leads to lower transformation rates (29, 30). These chemical agents solubilize the chloroethenes and inhibit reaction at the ZVI surface (30, 31). The effect of hydrophobic compounds is expected to be fundamentally different from surfactants and cosolvents, as the former specifically displace the chloroethenes from nonreactive sorption sites without directly interfering with the interaction of the chlorinated ethenes at the reaction sites. Although TCE sorption was approximately 4 times lower in the presence of the mixed BTX (Figure 2), this did not affect the TCE degradation kinetics as was observed for PCE (Figure 3). Although sorption of BTX to nonreactive sites would also make more TCE available for reaction, it should be noted that the extent of TCE sorption, even in the absence of competing solutes as BTX, was quite low (Table 1). This is illustrated by the fact that first-order kinetic parameters based on aqueous or total TCE concentrations were only about 20% different. The significant effect of BTX sorption on TCE sorption but not on TCE degradation again confirms the idea that sorption and reaction sites are fundamentally different entities (12). The results from the present study furthermore suggest that exposed graphite in ZVI did not mediate the reduction of chlorinated ethenes, as shown for the degradation of 2,4-dinitrotoluene by Oh et al. (13). Oh and co-workers (13) hypothesized that sorption of hydrophobic nonreducible co-contaminants (e.g., toluene) to

graphite sites would decrease overall reactions rates. However, in present study, an increased reduction of PCE and no effect on TCE degradation were observed in the presence of the BTX compounds (Figure 3). We did not observe significant effects of humic acids on the degradation of TCE, while PCE degradation kinetics decreased by 36% only in the presence of the higher humic acids concentration (100 mg L-1; Figure 3). Humic acid components, such as quinone moieties, have the ability to shuttle electrons from iron(III) oxides to iron-reducing bacteria and can act as electron acceptors for humicsreducing microorganisms (32). Reduced humic acids furthermore are able to directly reduce some chlorinated aliphatics (33). It might therefore be imaginable that HAs shuttle electrons from ZVI to chlorinated ethenes, potentially enhancing the dechlorination process (25). With a standard reduction potential E0 of 0.23 V (34), a model quinone as anthraquinone-2,6-disulfonate (AQDS) would theoretically be able to transfer electrons from ZVI (E0 ) -0.44 V (20)) for the reduction of PCE to TCE (E0 ) 0.58 V; (35)). In contrast, Tratnyek et al. (25) found that TCE degradation kinetics decreased by 21 and 39% in the presence of 20 and 40 mg L-1, respectively, of Suwannee River organic matter, as a result of competition for sorption to reactive surface sites. Although we did not observe any substantial competition for sorption between humic acids and chlorinated ethenes (Table 2), complexation of the humic acids by the oxide surface may have hindered the degradation of the less reactive species of PCE by blocking the reaction sites (Figure 3). Johnson et al. (2) found that both redox active and redox inactive organic ligands such as catechol, ascorbate, EDTA, and acetate decreased the carbon tetrachloride (CT) dechlorination rates through blocking of the interface sites where weaker interactions of CT with the iron oxide normally lead to dechlorination. It is not clear why TCE would be less impacted by these phenomena in our experiments. The landfill groundwater sample reduced the TCE and PCE degradation rates by 60% and 70% respectively (Figure 3). To our knowledge, there is only one other study examining the transformation of chlorinated organics in landfill leachate contaminated groundwater by ZVI (36). The authors found that the complex leachate matrix had little effect on the removal of PCE and DCE, in comparison to simple HEPESbuffered water. However, Schreier and Reinhard (36) observed that the reaction in buffered water ceased after about 42 d, while PCE and DCE transformation in the leachates stopped after 28 d. Moreover, biological reactions could not be excluded. The landfill groundwater sample used in our experiments was characterized by a negligible content of nitrate and chromate, which are known inhibitors of the ZVI process (37, 38), a relatively high specific conductivity (3150 µS cm-1, which was about 10 times the conductivity of the simulated groundwater), and a considerable amount of undefined dissolved organic carbon (DOC ) 109 mg of C L-1). It was shown previously that a 10-fold increase in the solution ionic strength did not affect the TCE degradation kinetics (38). It is therefore unlikely that the higher conductivity of the leachate caused the observed inhibition. The role of the organic carbon remains mostly unexplored. The nature of the organic carbon in landfill leachates depends on the “age” of the landfill (19, 39). An “old” landfill, like the source of our groundwater sample, is in the methanogenic phase of waste stabilization. In that case, the DOC content of the landfill consists mainly of refractory fulvic-like and humic-like compounds (19). Christensen et al. (40) characterized the DOC in landfill leachate polluted groundwater and found that a fulvic acid-like fraction (60%) and a hydrophilic fraction (30%) predominated. HA-like DOC made up only about 10% of the total organic carbon content. The probable presence of organic acids, acting as surface ligands, VOL. 38, NO. 10, 2004 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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in the groundwater DOC may explain the impact on the chloroethene degradation kinetics, as observed by Johnson et al. (2). The absorbance of the landfill groundwater sample at 254 nm, which is a measure for the dissolved organic carbon content, decreased by half in the presence of ZVI (results not shown). This indicates that at least part of the DOC interacted with the iron surface, potentially interfering with chloroethene reduction. Further characterization of the organic carbon present in the groundwater sample and elucidation of the role of the different organic fractions would be necessary to clarify our observations. Summarizing this study, we found that potential groundwater co-contaminants could impact the sorption and degradation of chlorinated ethenes in batch ZVI systems by competition for a finite number of reaction or sorption sites at the surface, depending on the physicochemical properties of the competitors. Notably, aromatic hydrocarbons displaced the chloroethenes from ZVI sorption sites, which resulted in higher PCE transformation rates. The significant effect of the complex landfill groundwater matrix on chloroethene reduction warrants further investigation, specifically with regard to the nature and role of the organic content.

Acknowledgments The authors would like to thank Miranda Maesen and Queenie Simons for their assistance during this work. This work was funded in part by a VITO Ph.D. grant and by EC Project QLK3-CT-2000-00163.

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Received for review August 26, 2003. Revised manuscript received March 7, 2004. Accepted March 8, 2004. ES034933H