Environ. Sci. Technol. 2008, 42, 5963–5970
Identifying Abiotic Chlorinated Ethene Degradation: Characteristic Isotope Patterns in Reaction Products with Nanoscale Zero-Valent Iron MARTIN ELSNER,* MICHELLE CHARTRAND,† NANCY VANSTONE, GEORGES LACRAMPE COULOUME, AND BARBARA SHERWOOD LOLLAR Stable Isotope Laboratory, University of Toronto, 22 Russell Street, Toronto, ON M5S 3B1, Canada
Received January 21, 2008. Revised manuscript received May 20, 2008. Accepted May 21, 2008.
Carbon isotope fractionation is of great interest in assessing chlorinated ethene transformation by nanoscale zero-valent iron at contaminated sites, particularly in distinguishing the effectiveness of an implemented abiotic degradation remediation scheme from intrinsic biotic degradation. Transformation of trichloroethylene (TCE), cis-dichloroethylene (cis-DCE), and vinyl chloride (VC) with two types of nanoscale iron materials showed different reactivity trends, but relatively consistent carbon isotope enrichment factors (ε) of -19.4‰ ( 1.8‰ (VC), -21.7‰ ( 1.8‰ (cis-DCE), and -23.5‰ ( 2.8‰ (TCE) with one type of iron (FeBH), and from -20.9‰ ( 1.1‰ to -26.5‰ ( 1.5‰ (TCE) with the other (FeH2). Products of the dichloroelimination pathway (ethene, ethane, and acetylene) were consistently 10‰ more isotopically depleted than those of the hydrogenolysis pathway (cis-DCE from TCE, VC from cisDCE), displaying a characteristic pattern that may serve as an indicator of abiotic dehalogenation reactions and as a diagnostic parameter for differentiating the effects of abiotic versus biotic degradation. In contrast, the product-related enrichment factors of each respective pathway varied significantly in different experiments. Because such variation would not be expected for independent pathways with constant kinetic isotope effects, our data give preliminary evidence that the two pathways may share an irreversible first reaction step with subsequent isotopically sensitive branching.
Introduction Metallic iron may effectively degrade reducible groundwater contaminants such as chlorinated hydrocarbons (1). Granular iron subsurface barriers are used to control and remediate contaminant plumes, and their short- and long-term performance has been investigated (2–4). Nanoparticulate iron is being proposed as a new remediation technology (5, 6). * Corresponding author phone: +49-89-31872565; fax: +49-8931873361; e-mail:
[email protected]. Present address: Institute of Groundwater Ecology, Helmholtz Zentrum Muenchen - National Research Center for Environmental Health, Ingolsta¨dter Landstraβe 1, D-85764 Neuherberg, Germany. † Current address: Earth Sciences Department, University of Ottawa, 140 Louis Pasteur, Ottawa, Ontario, Canada K1N 6N5. 10.1021/es8001986 CCC: $40.75
Published on Web 07/09/2008
2008 American Chemical Society
Slurries of fine iron particles (nanometer to micrometer scale) are injected into an aquifer where they are expected to spread over the area of highest contamination and facilitate in situ decontamination (7). While laboratory studies demonstrated the initial reactivity of such nanoparticulate Fe(0) (8–10), assessing field applications is more complex. Injected slurries may dilute and displace the contaminant plume so that decreasing compound concentrations alone are not necessarily proof of degradation effectiveness. Also, the resultant lower redox potential may stimulate microbial activity, leading to biodegradation as an alternative degradation route. Scheme 1 illustrates that the biodegradation of trichloroethylene (TCE) involves sequential hydrogenolysis, giving cisdichloroethylene (cis-DCE) and vinyl chloride (VC) as problematicintermediatesbeforecompletereductivedechlorination to ethene. Abiotic transformation, in contrast, is reported to involve mainly vicinal dichloroelimination, leading to completely dehalogenated acetylene, ethene, and ethane (11, 12). Recent publications suggested that compound-specific isotope analysis may help distinguishing between concurring pathways. Liang et al. measured more negative carbon isotopic enrichment factors in the abiotic reduction of perchloroethylene (PCE) and TCE on iron sulfide minerals compared to biodegradation (13). Our previous work with chlorinated ethanes showed significantly greater carbon isotope fractionation in β-dichloroelimination than in hydrogenolysis (14, 15). Since for chloroalkanes it is wellestablished that the initial reaction step is reductive bond cleavage (16), the results could be rationalized by a stepwise versus concerted character of initial C-Cl bond breakage. If this trend extended also to chlorinated ethenes, it would provide a novel approach to assess different degradation pathways at field sites, complementary to the analysis of compound concentrations. Earlier studies, however, reported more variable carbon isotope fractionation in abiotic chlorinated ethene reduction (17–20). Enrichment factors reported for degradation on zerovalent iron ranged from -5.7‰ (20) to -25.3‰ (17) for PCE, -8.6‰ (17) to -24.8‰ (18) for TCE, -6.9‰ to -16.0‰ (20) for cis-DCE, and -6.9‰ to -19.3‰ for VC (20). Notably, no systematic difference was found between PCE, TCE, and cisDCE, where β-dichloroelimination is expected to prevail, and VC, where only hydrogenolysis is possible. On the one hand, this may arise because abiotic dehalogenation of PCE, TCE, and cis-DCE does not occur exclusively by β-dichloroelimination, but also involves some hydrogenolysis. Estimates range from 15 to 20% hydrogenolysis depending on the type of iron or iron mineral (21, 22). On the other hand, chlorinated ethenes contain a double bond so that a wider range of reaction mechanisms is possible, for example, the postulated initial formation of di-σ-bonded surface complexes (12). Scheme 2 illustrates that, because different scenarios may be considered, interpretation is more complex than with chlorinated ethanes. Scenario A. Chlorinated ethene reduction proceeds by two distinct pathways such that reductive β elimination and hydrogenolysis do not share a common intermediate. Each pathway has its constant kinetic isotope effect. Consequently, the isotopic discrimination between the instantaneously formed respective product and the substrate (i.e., the productrelated ε) is not affected by the relative importance of hydrogenolysis versus β elimination. In contrast, enrichment factors determined from substrate data according to the Rayleigh model represent a weighted average of the two pathways occurring simultaneously and are strongly dependent on how much each pathway contributes. VOL. 42, NO. 16, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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SCHEME 1. Concurring Pathways Proposed in Chlorinated Ethene Dehalogenationa
a Biodegradation generally involves sequential hydrogenolysis to ethene (upper pathway), whereas abiotic dehalogenation includes also vicinal dichloroelimination (lower pathway).
SCHEME 2. Different Mechanistic Scenarios for Abiotic Chlorinated Ethene Reduction (Example cis-DCE)a
a ε is the enrichment factor of the Rayleigh equation, ε cis-DCE->dichloroelimination the enrichment factor of the dichloroelimination and εcis-DCE->hydrogenolysis that of the hydrogenolysis product formation. In scenario B, ε1 and ε2 are enrichment factors of subsequent reactions after the first irreversible step.
Scenario B. Both pathways involve a common irreversible step. The ε value derived from the Rayleigh equation depends only on the kinetic isotope effect of this step and hence is constant and independent of the dominant pathway. The hydrogenolysis product may nonetheless be isotopically different from the β elimination product, since different bonds are broken or formed in consecutive reactions of the common intermediate. The isotopic discrimination between the two products is constant, because it depends only on the constant kinetic isotope effects of these subsequent reactions, denoted by ε1 and ε2. The isotopic discrimination between the substrate and each product (i.e., the product-related ε), however, is no longer constant but, for reasons of isotopic mass balance, varies as a function of respective product yield, as reviewed by Hayes (23). In this study, we investigated carbon isotope fractionation manifested in the substrate and products of TCE, cis-DCE, and VC transformation with nanoscale zero-valent iron. Two different iron types were chosen that represent distinct methods of synthesis, have been characterized in previous studies (8–10, 24), and are the most common types used commercially for remediation in the field. One (FeBH) was 5964
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obtained from Golder Associates and is produced by a reduction of dissolved FeII through NaBH4 (8). The other (FeH2) was purchased from Toda Kogyo Corp. and is manufactured by a reduction of goethite and hematite particles with H2 gas at 200-600 °C. Our objectives were (i) to investigate the consistency of carbon isotopic enrichment factors (εC) for different chlorinated ethenes with the different types of nanoirons and compare them to values reported for granular Fe(0) and FeS, (ii) to determine reaction kinetics and products to assess the relative contribution of concurring pathways, (iii) to measure isotope discrimination in the reaction products and determine whether independent pathways prevail (scenario A) or whether a common intermediate is formed (scenario B), and finally, (iv) to assess the potential of such isotope patterns to evaluate the effectiveness of nanoscale zero-valent iron applications at contaminated sites.
Materials and Methods Experimental Setup. Details about chemicals, experiments, and analytics are provided in the Supporting Information (SI). Nanoparticulate iron from Toda Kogyo Corp. (“FeH2”)
and from W.-X. Zhang/Golder Associates (“FeBH”) was received as a suspension in water. Properties of the materials are described in ref 24. About 6 months after production, the iron materials were freeze-dried under anoxic conditions according to ref 25 and resuspended in anoxic TRIS buffer at pH 7 to give 25 g/L of FeBH and 9 g/L of FeH2, respectively. While not further characterized, they represent therefore aged rather than freshly synthesized material. Experimental vials contained 100 mL of aqueous suspension and 150 mL of headspace and were equipped with a Mininert valve for sampling. They were constructed under anoxic conditions (95% N2, 5% H2) in an anoxic chamber, taken out, and rotated on their side at 45 rpm to ensure rapid solid/water and water/ air mass transfer. Concentration and Isotope Analysis. After direct headspace sampling, concentrations of organohalide reactants and products were determined on a gas chromatograph equipped with a flame ionization detector using a 90 m GS-Q column. Isotope ratios for all compounds were analyzed by direct headspace analysis on a gas chromatograph-combustion-isotope ratio mass spectrometer (Finnigan MAT 252) and are reported as δ13C values relative to the international standard V-PDB. δ13C ) (R/RVPDB -1) × 1000, where R and RVPDB are the 13C/12C ratios of the compound and the international standard, respectively. Further details are provided in the SI. Evaluation of Carbon Isotope Fractionation. Substrate isotope data were evaluated according to a Rayleigh regression not forced through the origin (26): (1000 + δ13C) R ) ) f (εC⁄1000) R0 (1000 + δ13C )
(1)
of a particular pathway. It is well-established that the weighted average isotope value of all products δ13Cproduct,average shows a trend according to (28, 29) 1000 + δ13Cproduct,average 13
1000 + δ C0,substrate
1 - f (εC⁄1000+1) 1-f
)
We adopted this equation for the case where several products are formed in parallel and their respective isotope ratios deviate from the weighted average, such as in scenario A or B (Scheme 2). Because kinetic isotope effects of parallel pathways do not change during a reaction, the isotopic difference between parallel products and their weighted average also remains constant (14). Consequently, we derived the general fitting equation: 1000 + δ13Cproduct 1000 + δ13C0,substrate
) (1 +
D(δ13C) 1 - f (εC⁄1000+1) ) 1000 1-f
εSubstratefProduct ) δ13C0,product - δ13C0,substrate ) [D(δ13C) + δ13C0,product,average] δ13C0,substrate 13
εSubstratefProduct Rproduct,instantaneous -1 ) 1000 Rsubstrate )
δ13Cproduct,instantaneous - δ13Csubstrate
(2)
For their determination, two strategies were employed: (I) During the first 5-10% of conversion, a closed reaction vial behaves approximately like an open system where accumulation is negligible and this instantaneous discrimination is directly observable (27): εSubstratefProduct ≈ δ13C0,product - δ13C0,substrate
(6)
13
) D(δ C) + [δ C0,product,average δ13C0,substrate] 13
) D(δ C) + εC is calculated from the parameters D(δ13C) and εC. Provided that the initial turnover of intermediates can be estimated, eqs 3 and 6 allow the determination of product-related fractionation even without a knowledge of absolute reaction rates, product distribution, or closed molar balances, since they rely solely on isotope measurements of the substrate and a given product.
Results and Discussion
(1000 + δ13Csubstrate)
δ13Cproduct,instantaneous - δ13Csubstrate ≈ 1000
(5)
where D(δ13C) ) δ13Cproduct - δ13Cproduct,average expresses by how much a particular product isotope value deviates from the weighted average of all products (14). Note that the detection of all products or closed molar balances is not required here, since eq 1 provides the enrichment factor of the weighted average, εC, and eq 5 provides the deviation from this average for a particular product, D(δ13C). Subsequently, the product-related isotopic discrimination
0
where R0 and R are carbon isotope ratios at the beginning and a given time (t) during reaction, δ13C0 and δ13C are the same values in delta notation, f is the fraction of substrate remaining at time t, and C is the isotopic enrichment factor. Product-related enrichment factors εsubstratefproduct are defined by the discrimination between isotope ratios of a given substrate Rsubstrate and each of its instantaneously formed products Rproduct,instantaneous:
(4)
(3)
with δ13C0,substrate and δ13C0,product being the initial substrate and product isotope values, respectively. If the instantaneous product is further transformed, two cases may be distinguished (see ref 23, Figure 3). If the product’s turnover is significantly smaller than its formation, its isotope value still approximately reflects the instantaneous formation. In contrast, if the consecutive reaction is significantly faster, εsubstratefproduct must be derived from the weighted isotopic average of all products of the pathway. (II) Alternatively, if the consecutive reaction is much faster and intermediates do not accumulate significantly, observable end products directly reflect the isotopic discrimination
Reactivity Trends. Table 1 summarizes reactivity trends with the different nanoscale iron materials when TCE, cis-DCE, and VC were separately brought to reaction. Suspensions of 9 g/L of FeH2 achieved a complete transformation of TCE to dehalogenated products in 10 days (Figure 1). Under identical experimental conditions, the conversion of cis-DCE was less than 5%, and the transformation of VC was negligible (Table 1). The opposite trend was observed for suspensions of 25 g/L of FeBH, where reaction rates were faster for lesschlorinated substrates (kobs for VC > cis-DCE > TCE; Figure S1). With FeH2, the reactivity trend therefore corresponded to expectations from redox potentials (30), whereas with FeBH, our results are similar to those of Arnold and Roberts, who postulated di-σ-bonded surface intermediates (12). It remains to be investigated whether the different trends observed in this study indeed give evidence of different initial reaction mechanisms (electron transfer versus organometallic complex formation). Consistent with previous results (10), FeBH showed lower reaction rates for higher concentrations of chlorinated ethenes, indicating competition for available surface sites (Table 1). In contrast, FeH2 displayed slightly faster rates at higher TCE concentrations, suggesting that surface reactivity was regenerated during reaction. Possibly, such an effect VOL. 42, NO. 16, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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-23.1 ( 4.8 -20.1 ( 2.8 -19.0 ( 1.3 -9.5 ( 1.5
( ( ( (
1.0 0.93 0.52 0.81
acetylene
0.01 0.04 0.01 0.09
-21.8 ( 2.2 -22.5 ( 1.4
0.13 ( 0.01 0.32 ( 0.02
2.8 1.2
cis-DCE
0.18 0.59 0.76 1.46
-23.1 ( 5.3 -23.7 ( 3.1 -21.4 ( 2.9
0.08 ( 0.01 0.14 ( 0.08 0.12 ( 0.01
0.56 0.55 3.2
VC acetylene TCE
VC
n.q.g n.q.g n.q.g
