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cesses even more difficult. New strategies are, therefore, necessary that enable a better qualitative and quantitative assessment of transformation processes.
Assessing Transformation Processes of Organic Compounds Using Stable Isotope Fractionation
Determination of compound-specific stable isotope signatures
THOMAS B. HOFSTETTER,* ´ P. SCHWARZENBACH, AND RENE STEFANO M. BERNASCONI Swiss Federal Institute of Technology (ETH) Zurich
ISTOCKPHOTO/THOMAS HOFSTETTER/RHONDA SAUNDERS
Compound-specific stable isotope analysis makes it possible to infer the origin and transformation pathways of organic compounds.
Compound-specific stable isotope analysis (CSIA) at natural abundance isotope levels opens new avenues to assess the sources and transformation processes of organic compounds and is, in principle, applicable to any isotopic element in an organic compound. Using gas and liquid chromatography coupled to isotope ratio mass spectrometry (1–3), isotope ratios of different elements can be measured in individual organic compounds. To date, the most commonly analyzed elements are carbon (i.e., molar ratios of 13C/12C) and hydrogen (2H/1H) and, to a much lesser extent, nitrogen (15N/ 14N), oxygen (18O/16O), and chlorine (37Cl/35Cl). Since the abundance of rare isotopes and the variations in isotope ratios of these elements are generally very small, isotope ratios are given in the “delta notation”, δhE, in “per mil” relative to internationally defined standard materials (hE/ lE reference, Equation 1). δhE )
(
h
E/1Esample
h
E⁄ 1Ereference
)
- 1 · 1000
(1)
For example, 13C/12C ratios are typically in the range of 0.0112, and variations as little as 0.000006, corresponding to a shift in δ13C of (0.5‰, can be quantified precisely in routine analysis (4). The more abundant the heavy isotope ratio, the better the analytical precision typically is. (5; Box 1).
What type of information can be obtained from isotope signatures?
Assessing transformation processes of organic compounds in the environment is crucial for addressing the risks of soil and water contamination and for estimating substance fluxes in biogeochemical cycles. Quantifying such processes by conventional means is usually a major challenge. The measurement of concentration changes of organic compounds in environmental samples often does not allow unambiguous detection of transformations, due to effects of dilution, volatilization, or sorption to the environmental matrix. Moreover, concentration measurements in soils and aquifers are analytically challenging and require the installation and operation of expensive and labor-intensive sampling networks. Finally, transformations of organic compounds can proceed simultaneously along competing pathways, thus making the characterization of these pro10.1021/es801384j
2008 American Chemical Society
Published on Web 10/30/2008
Variations in the isotope composition of organic compounds exhibit some key features that make them invaluable for allocating sources and for tracing (bio)chemical transformations. As a first approximation, it is usually assumed that only the cleavage or formation of a chemical bond generates significant shifts of a compound’s isotope signature. It is irrelevant whether the reaction takes place abiotically (e.g., at a mineral surface) or is enzyme-catalyzed, as long as the bond cleavage or formation is at least partly limiting the overall rate of reaction. In contrast, phase transfer (e.g., sorption, volatilization) or transport processes (diffusion) generally lead to much smaller changes of isotope compositions. Organic compounds have a characteristic “isotopic fingerprint”, defined by their precursor materials’ isotope composition and by the compounds’ formation process(es). If the concentration of a compound in the environment only changes owing to dilution or phase transfer processes, its isotopic composition will remain largely unaffected and can thus be used to elucidate its sources. To determine whether organic chemicals are naturally occurring or man-made, radiocarbon contents VOL. 42, NO. 21, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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Box 1 Determination of compound-specific stable isotope signatures in organic compounds Online methods for measuring isotopic compositions in organic compounds differ primarily by the conversion of the target molecules to analyte gases in interfaces between chromatographic separation and the isotope ratio mass spectrometer (2). The table below lists the isotopic elements that can be analyzed using the most widely applied, commercially available gas chromatographs/isotope ratio mass spectrometers (GC/IRMS). Depending on the considered element, interface processes include combustion (13C), combustion/reduction (15N), and pyrolysis (2H, 18O). Isotope ratio 2H/1H 13C/12C 15N/14N 18O/16O a
Abundance in standard(%)a
Reference materialb
Analyte gas
Typical precision (‰)a
Typical sensitivity (nmol)a
0.0115 1.12 0.366 0.201
Water (VSMOW) Carbonate (VPDB) Air (AIR) Water (VSMOW)
H2 CO2 N2 CO
2-5 0.1-0.3 0.3-0.7 0.3-0.6
10-50 0.1-5 1-10 4-14
Data modified from ref 2.
b
See ref 31 for details.
Alternative avenues to stable isotope analyses in organic compounds are being developed. Liquid chromatography can be coupled to IRMS via wet oxidation or moving-wire interfaces. Owing to issues with eluent removal, these techniques are at present limited to 13C analysis (3). Competing approaches for compound-specific Cl and Br isotope analysis have recently been reported and include GC/IRMS analysis without interface (currently restricted to chlorinated ethenes; (32) or GC coupled to multicollector ICP-MS (33, 34). For sulfur isotope analysis in organic compounds, no online methods are yet available.
(δ14C) are frequently used in addition to stable isotopes. Biologically synthesized organic compounds ultimately derive their carbon from atmospheric CO2, which contains 14C formed from interaction of 14N with cosmic rays, whereas industrial products synthesized from petroleum products are considered “14C-free” because 14C has decayed completely. This allows one to distinguish, for example, natural bioaccumulative polybrominated biphenyl ethers formed via biosynthesis from precursors, which exhibit high 14C contents, from petroleum-derived industrial products (6). Similarly, the formation of polycyclic aromatic hydrocarbons can be apportioned to their contributions of fossil fuels vs biofuels (7). Finally, differences in isotope compositions help to identify the producer or sources of organic pollutants. For example, production of chlorinated ethenes by different manufacturers can generate subtle differences in δ13C and δ37Cl during industrial synthesis (8). In the environment, temporal and spatial variations in isotope signatures of organic compounds can reveal if and by which pathway transformations take place. The extent and direction of isotope fractionation provide clues for identifying the reactions and mechanisms leading to degradation or formation. If quantitative information on the fractionation behavior is available, isotope signatures may even reveal to which degree a reaction has progressed.
Assessing transformation pathways Contaminant degradation. An increasing number of studies illustrate how CSIA can be applied for the quantitative degradation assessment and for the elucidation of transformation mechanisms of various compound classes. In the following, examples are used to illustrate how changes in isotope signatures provide unique information on reaction pathways. First we consider the biodegradation of nitroaromatic compounds (NACs), a major class of soil and groundwater contaminants owing to their widespread use as pesticides, dyes, explosives, and industrial feedstocks. Nitroaromatic compound biodegradation occurs along different, sometimes competing, enzymatic and abiotic reaction pathways, depending on their molecular structure and environmental conditions (9). Biodegradation of nitrobenzene in aerobic soils, for example, can proceed simultaneously via oxidation to catechol or via partial reduction to an aminophenol. In such cases, evidence for the predominant degradation pathway and estimates of the extent of bio7738
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degradation are difficult to obtain from analysis of reactant concentrations alone. An evaluation of reaction product concentrations also may not be conclusive: some products may already be present in the soil matrix (e.g., nitrite or catechols from nitrobenzene oxidation), or they can be further degraded by other microorganisms. Combined analysis of stable carbon and nitrogen isotope fractionation, in contrast, facilitates both an identification and quantification of the prevailing type of reaction (10). Figure 1 illustrates how the δ15N, δ13C, and δ2H values of residual nitrobenzene increase during progressive biodegradation by Comamonas strain JS765. The degree of enrichment can be quantified with bulk isotope enrichment factors, εE. They exhibit the more negative values the more preferred is the reaction of the isotopically lighter reactant is (in this example: nitrobenzene containing 2H, 12C, 14N). Equation 2, which can be used to calculate either curve in Figure 1 (a modified form of the Rayleigh distillation equation), shows that the isotopic enrichment depends on the fractional conversion of the substrate but is independent of the rate of the process:
(
δhE + 1000 c ) h c0 δ E0 + 1000
)
1000 ⁄εE
(2)
where c and c0 are the compound’s current and initial concentration and δhE and δhE0 are its current and initial isotope signatures, respectively. Isotope enrichment factors are characteristic for a given degradation pathway because, depending on the underlying reaction mechanism, isotope fractionation occurs according to a kinetic isotope effect (KIEE) per element E at the reacting bond(s). KIEEs reflect the different reaction rates of bonds containing light or heavy isotopes (i.e., KIEE ) lk/hk) and result from the difference in energy of activation between the ground state of the isotopic reactants and their respective transition state in the reaction (11). This is in contrast to equilibrium isotope effects, EIEEs, which are ratios of KIEEs for the forward and backward reaction, respectively, and only depend on bonding changes between reactant and product. Kinetic isotope effects are large if bonds are broken or formed (primary KIEEs) in the rate-limiting step. This is illustrated by the aerobic nitrobenzene biodegradation (Figure 1). During the dioxygenation of nitrobenzene to a tentative intermediate compound 1, two of six C atoms change their
FIGURE 1. Hydrogen, carbon, and nitrogen isotope fractionation during the oxidative biodegradation of nitrobenzene by Comamonas sp. strain JS765 (10). (a) Isotope enrichment measured in the bulk reactant during the transformation process, c/c0 stands for the fraction of remaining nitrobenzene. Lines are calculated using Equation 2. (b) Linearized isotope enrichment behavior used for the derivation of H, C, and N isotope enrichment factors εE from the linear regression slopes according to Equation 3 (Box 2). AKIEE stands for the apparent kinetic isotope effects derived from the observed εE values (see Box 2 and Equation 5). bonding from CdC double bonds to C-C single bonds and form new bonds to oxygen with an apparent KIEC of 1.024
(see Box 2 for details on calculation of apparent KIEEs). 12Ccontaining bonds thus react 2.4% faster than 13C-containing bonds. As no chemical bond to N in the NO2 group is broken at this stage, the 15N fractionation is small (secondary apparent KIEN of 1.0008, Figure 1). For the same reasons, the corresponding H isotope fractionation (KIEH of 1.045) is also a secondary one as primary H isotope effects are much larger (generally 2-8; 12). The magnitude of kinetic isotope effects cannot be predicted from the molecular structure of reactant in the absence of information on transition state structures. Even though reference experiments from the (bio)chemical literature are available for comparison, mechanistic studies on isotope effects in environmental reactions are scarce. In the absence of experiments, some rules of thumb can be used to address qualitative trends. For example, KIEEs increase with increasing relative masses of the isotopes (e.g., 2H/1H > 13C/12C). Moreover, assuming similar bond strength, the KIEE of element E increases with increasing mass of the bonding partner. For example, 13C kinetic isotope effects increase in the sequence C-H < C-C < C-Cl (see further details in ref12). Distinguishing competing transformation pathways by multielement isotope analysis. Whereas different reactions can exhibit the same isotope fractionation for one element and εE-values overlap within experimental error, it is much less likely that this happens for several elements simultaneously. A multidimensional isotope analysis is thus key for identifying the predominant degradation pathways. For example, the competing oxidative and reductive biodegradation pathways of nitrobenzene under oxic conditions are characterized by a significant 13C and small 15N fractionation for the oxidation to catechol, while the partial reduction to o-aminophenol shows the opposite δ13C and δ15N trends (Figure 2a). During biodegradation, the isotopic composition of residual nitrobenzene in a δ13C vs δ15N diagram will thus evolve with a small slope during oxidation compared with a steep slope during reduction according to each pathway’s ratio of εN/εC. Even if both reactions happen simultaneously, the relative share of either pathway is revealed by the trends of the reactant’s isotope signature change, as shown by the solid lines in Figure 2a. Note that this approach is independent of the measured reactant concentrations, a fact that greatly simplifies the applicability of multidimensional isotope analysis in the environment (13). Similar multidimensional isotope analyses have been performed for other compound classes such as the fuel constituents methyl-tert-butyl ether (14), BTEX compounds (15, 16), chlorohydrocarbons (17), and triazine herbicides (18). Quantifying the extent of transformation. As outlined above, changes in isotope signatures are a function of the extent to which a compound has been transformed. The larger the difference between the original compound’s δhE and its value at a later stage of the reaction, for example, downstream of a contamination source, the more it has been transformed. Such a quantitative assessment is possible with a single enrichment factor once the predominant transformation process has been identified (see Equation 2) and as long as the reaction kinetics are determined by the same elementary step. This is, unfortunately, not always the case, as illustrated for BTEX degradation in anoxic, fuel-contaminated subsurface environments. Under such conditions, the coupling of microbial oxidation of BTEX compounds to dissimilatory iron reduction is an important contaminant degradation process. The oxidation of toluene, which is initiated via cleavage of a C-H bond at the methyl group, shows different isotope enrichment factors for C and H (εC and εH) depending on the availability of Fe(III) as terminal electron acceptor (19). Isotope fractionation in solutions VOL. 42, NO. 21, 2008 / ENVIRONMENTAL SCIENCE & TECHNOLOGY
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Box 2 Determining isotope enrichment factors and calculating apparent kinetic isotope effects Isotope enrichment factors, εE, are calculated assuming that the observed isotope fractionation in the bulk organic compound is caused by intermolecular isotopic competition between molecules containing exclusively n isotopically light atoms of element E (i.e., the light isotopologue) vs isotopologues containing one heavy and (n - 1) light isotopes (11). For example, 13C fractionation during chloroethene transformation is due to competition between isotopologues containing two 12C atoms vs isotopologues containing one 12C and one 13C atom. The concentration of isotopologues with two 13C atoms is, according to a binomial distribution of light and heavy isotopes, too small to contribute to the measurable isotope fractionation. εE values are obtained from linear regressions of measured δhE values and the fraction of remaining substrate (as normalized concentrations c/c0).
(
ln
δhE0 + ∆δhE + 1000 δhE0 + 1000
)
)
()
εE c · ln 1000 c0
(3)
where δhE0 is the initial isotope signature and ∆δhE is its change relative to the initial value. While bulk εE values are used to assess the fractionation of entire molecules, this parameter is not suited to comparing isotope effects and reaction mechanisms among different compounds. For the latter, kinetic isotope effects at the reacting bond are necessary. To convert εE values into apparent kinetic isotope effect, εEs are calculated as isotope enrichment factors at the reactive position of the molecule, εreactive position.
(
ln
δhE0 + n ⁄ x · ∆δhE + 1000 h
δ E0 + 1000
)
)
()
εreactive position c · ln 1000 c0
(4)
The correction for n isotopic atoms of element E and x reactive sites within the molecule takes into account that (i) nonreactive atoms ”dilute” the observed isotope fractionation whereas (ii) multiple reactive sites can increase it. Finally, intramolecular isotopic competition can lead to a depletion of heavy isotope containing molecules without leading to isotope fractionation (see examples in Box 3). This is the case if a molecule reacts at the isotopically light reactive site instead of the one carrying the heavy isotopes. Correction for z competing reactive sites leads to the apparent kinetic isotope effect, AKIEE (for details see Ref. 12). 1 (5) AKIEE ) 1 + z · εreactive position ⁄ 1000 containing dissolved Fe(III) was almost 3 times larger than if bacteria used a solid Fe(III) mineral for respiration (Figure 2b and Box 3). A calculation with Equation 2 for an arbitrary shift of ∆δ13C of toluene of 1‰ using εC values for toluene oxidation in solution versus εC values for suspensions results in very different estimates for the extent of transformation (24% with dissolved Fe(III) vs 54% if solid mineral Fe(III) is the electron acceptor). Apparently, toluene oxidation rates in suspension can be limited by the transport of toluene to the bacteria colonizing mineral particles. This transport process thus masked the isotope fractionation associated with the bond-cleaving reaction in suspensions, while no masking occurs in homogeneous solution. Two-dimensional isotope analysis reveals the same degradation pathway regardless of masking phenomena. Even if an isotope fractionating step is not fully rate-limiting, relative isotope signature changes in the case of toluene oxidation ∆δ2H/ ∆δ13C, are identical to those under nonmasked conditions (Figure 2b). However, multidimensional approaches cannot circumvent the problem that variable enrichment factors for the same reaction render it difficult to calculate unequivocally the fraction of (bio)degraded contaminant (13). Elucidating process networks. Isotope fractionation not only reveals degradation pathways but also allows inference of the origin and formation pathways of products even if they are generated by a sequence of reactions. Such information is essential for identifying material fluxes and processes driving the biogeochemical cycling of carbon (20). Because isotope enrichment in a reactant is coupled to a corresponding isotopic depletion in the product (and vice versa), products can be related to their precursors. For example, δ2H and δ13C values of methane differ quite significantly depending on whether methanogenesis occurred via CO2 reduction, by fermentation of acetate or formate, or whether methane was produced by thermal cracking of sedimentary organic matter or petroleum (21). 7740
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In addition to the two-dimensional evaluation of contaminant transformation reactions (Figures 2a and 2b), changes in ∆δ2H/∆δ13C of methane can be used to detect variations in the methane’s substrate pools vs consumption by methane oxidation. The approaches used to interpret variations in isotope signatures of biologically synthesized substances are similar to those used for transformation processes of single organic compounds. However, for the former, strong variations in isotope enrichment factors are often observed for apparently similar reactions, because most biogeochemical processes are composed of multiple steps in complex process networks. The overall isotope fractionation between source and final product is influenced not only by the rate-limiting reaction steps but also by branching points in the reaction chain and by the removal of intermediate products from the considered system (20). While an identification of the crucial transformation processes is possible despite these variations, a reliable quantitative interpretation of compound fluxes via isotope fractionation can only be achieved in relatively simple systems.
Elucidating reaction mechanisms Assessing transformation processes from observed isotope fractionation is most reliable if bulk isotope enrichment factors can be interpreted in terms of bondspecific, apparent KIEEs (Boxes 2 and 3). This requires knowledge of the underlying reaction mechanisms and the kinetics of the elementary reactions up to an irreversible bond cleavage (22). Unfortunately, for most environmental reactions, including enzyme- or surface-catalyzed redox reactions, phototransformations, and so on, neither are “typical” ranges of isotope effects known nor are the process kinetics elucidated to a point where rate-limiting steps have been identified unequivocally. Recent studies have, therefore, reevaluated various “established” degra-
first N-O bond after sequential transfer of two electrons and two protons. At low redox potentials and under protonlimited conditions, however, this transformation proceeds along a different mechanism leading to the same observable products, where electron transfers become ratelimiting thus reducing apparent 15N kinetic isotope effects by up to a factor of 4 (24). Such effects have not yet been confirmed experimentally, but computational results support current hypotheses. Dealing with such variations in isotope effects and thus isotope enrichment factors can be a major drawback for the application of CSIA. Such problems can be circumvented with mechanistic studies, which show how and why environmental conditions modulate isotope fractionation.
Assessing transformation processes in the field
FIGURE 2. Two-dimensional isotope analysis for the evaluation of contaminant degradation pathways. (a) Changes of δ15N and δ13C values of nitrobenzene during its biodegradation via oxidation to catechol (red symbols) vs partial reduction to aminophenol (blue symbols) (reproduced with permission from Ref 10). Shaded areas indicate exclusive occurrence of either one of two pathways (95% confidence intervals). Solid lines represent the evolution of isotope signatures if both processes occur simultaneously in arbitrary ratios of 3:1, 1:1, and 1:3 (as relative shares of reduction and oxidation). Dashed lines show the extent of nitrobenzene biodegradation based on nitrobenzene’s initial signatures. (b) Changes of δ2H and δ13C values of toluene during its anaerobic oxidation by a dissimilatory iron-reducing microorganism with dissolved Fe(III) (black symbols/lines) vs a solid Fe(III) mineral as terminal electron acceptor (red symbols/lines, Ref. 19). Molecular structures illustrate the isotopic reaction step, that is, methyl group oxidation, regardless of the type of Fe(III) used by the bacteria. dation reactions such as abiotic reductions of nitroaromatic contaminants, photolysis, and substitution reactions of triazine herbicides, to fill some mechanistic gaps (18, 23). In this context, computational chemistry is used to address possible isotopic reaction steps of intermediates that cannot be accessed experimentally. Moreover, transition state structures can be postulated to help rationalize apparent KIEEs. Successful computations finally can provide some benchmark values of isotope effects to which experiments can be compared. For example, the abiotic reduction of nitroaromatic compounds at surfaces of Fe(II)bearing minerals and the accompanying N isotope fractionation are usually determined by the cleavage of the
Contaminant transformation studies at complex field sites illustrate the potential of CSIA for identifying such processes successfully, especially when multiple elements are evaluated simultaneously and information is interpreted in terms of reactive transport models. As shown for MTBE (14) or BTEX transformation (25), combined δ2H and δ13C analysis of field and laboratory model systems revealed biotransformation and allowed inference or confirmation of the predominant reaction mechanism(s). Additional evidence for biotransformation can also be found via isotopic analysis of terminal electron acceptors, which are often difficult to link directly to the fate of individual organic compounds. In combination with CSIA of BTEX compounds, for example, sulfur and oxygen isotope fractionation in sulfate pointed toward a contribution of sulfate-reducing bacteria to contaminant transformation (26). Given the detailed analysis of the microbial sulfate reduction network and the associated isotope fractionation steps, assessing the reaction progress and even transformation rates may be possible using sulfate’s sulfur and oxygen isotope signatures (27). Reductive dechlorination of chlorinated hydrocarbons such as tetrachloroethene (PCE), trichloroethene (TCE), or 1,1,2,2-tetrachloroethane at solvent-contaminated sites also highlights some future research opportunities. Owing to the analytical difficulties of Cl isotope analysis, chlorohydrocarbon transformation assessment by CSIA has, to date, been largely limited to C isotopes. Because the mechanisms by which dehalo-respiring microorganisms reduce various chloroethenes are not fully understood, the significant variations of 13C enrichment factors reported for different organisms and different compounds are difficult to rationalize (28–30). In addition, carbon isotope effects upon biodegradation can be masked by the dissolution kinetics if chlorohydrocarbons are present in nonaqueous phase liquids or by diffusion in the gas phase of the unsaturated zone of aquifers. Consequently, quantitative assessment of chlorohydrocarbon biodegradation can be accompanied by larger uncertainties unless product isotope ratios can be used to constrain the observed isotope fractionation. Recent analytical advances in halogen isotope analysis (see Box 1) suggest that multidimensional isotope analysis will be available in the near future to address some of these problems. Thomas B. Hofstetter is a lecturer and senior scientist at the Institute of Biogeochemistry and Pollutant Dynamics, Stefano M. Bernasconi is a lecturer and senior scientist at the Geological Institute, and Rene´ P. Schwarzenbach is professor of environmental chemistry and head of the department of environmental sciences. All authors are affiliated with the Swiss Federal Institute of Technology (ETH) Zurich. Address correspondence about this article to Hofstetter at thomas.
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Box 3 From bulk isotope enrichment factors to bond-specific apparent kinetic isotope effects The procedure outlined in Box 2 can be applied to quantify the typical isotope fractionation behavior at the reacting bonds of organic compounds. The resulting apparent kinetic isotope effects, AKIEEs, are characteristic for underlying reaction mechanism (12). Broken bond Elements
Observed εE Number of atoms Reactive sites εreactive position Competing sites Observed range of AKIEE (-) values (‰) (n) (x) (‰) (z) AKIEE (-)
Oxidation of a C-H bond. Example for anaerobic enzymatic oxidation of toluene coupled to the microbial reduction of a solid iron mineral (19, 35) C-H C -1.3 7 1 -9.3 1 1.0093 1.005-1.02 H -35 8 3 -83 3 1.35 1.3-3.0 Reduction of a N-O bond. Example for reduction of 4-nitrotoluene by mineral-bound Fe(II) species (23, 36) N-O N -39 1 1 -39 1 1.040 1.030-1.045 Reductive dechlorination of a C-Cl bond. Example for pentachloroethane reduction to trichloroethene (17) C-Cl C -15 2 1 -30 1 1.030 1.02-1.03 Cl no data available 1.021 unknown
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