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5 Relating Isotope Fractionation to Isotope Effects and Reaction Mechanisms

5 Relating Isotope Fractionation to Isotope Effects and Reaction Mechanisms

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D. Hunkeler



598



where

Lk is

2



the rate constant for the reactive step when a light isotope is present at

the reacting site

Hk is the rate constant for the reactive step when a heavy isotope is present at

2

the reacting site

This difference in the rate constant of the reactive step is the fundamental cause

of isotope fractionation and therefore often denoted as the intrinsic KIE. The

apparent kinetic isotope effect (AKIE) reflects differences in rates that are detected

when the substrate A outside the cell is considered, and thus might be altered by

rate-limiting steps preceding the reaction. The AKIE can be expressed as follows

(Cleland 2006; Hunkeler and Elsner 2010):



AKIE =



Lk

2

Hk

2



+C



1+C



=



KIE + C

1+C



(25.20)



where C is the commitment to catalysis, which reflects the tendency of the

enzyme-substrate complex to undergo reaction versus dissociation



C=



Lk

2



k−1



(25.21)



A high C corresponds to the situation where each molecule that binds reacts

immediately independent of its isotope substitution. In this case, AKIE converges

towards 1 and no isotope fractionation is detectable because the molecules do not

get back into solution where measurement takes place. In reverse, if C is small,

the full KIE is detectable because the molecules that have undergone isotope fractionation partition back into the bulk solution. This effect of steps preceding the

reaction is one of the reasons why isotope fractionation factors can vary somewhat

even if the compound is degraded by the same mechanisms having the same KIE.

The second effect (“dilution”) occurs because KIEs reflect the presence of a

heavy isotope at a specific position in the molecule while CSIA and isotope fractionation factors represent the average behaviour of a heavy isotope located at any

position in the molecule. Therefore, if a heavy isotope is present at a non-reacting

position it would not influence the reaction rate but it would damp the observed

isotope effect. Two situations can be distinguished. A molecule can have several

identical reactive positions such as the two C–Cl bonds in 1,2-DCA that compete

for reaction (intramolecular competition). For example, during nucleophilic substitution of 1,2-DCA, a heavy chlorine isotope might be present in the reacting

bond causing primary isotope effect, or in the adjacent position having little effect.

Furthermore, the heavy isotopes might be present at positions that never react. For

example, in vinyl chloride some of the heavy carbon atoms will be at the position

without chlorine and hence will never participate in reactive dechlorination. The

AKIE can be related to the isotope fractionation factor using the following equation (Elsner et al. 2005):



25  Use of Compound-Specific Isotope Analysis (CSIA) …



AKIE =



1+z·



n

x



1

· (α − 1)



599



(25.22)



where

n

is the number of atoms of the considered element that are present in the

molecule

xis the number of atoms located at reactive sites

z

is the number of atoms located at reactive sites that are in intramolecular

competition

Using carbon isotope fractionation during degradation of 1,2-DCA by nucleophilic substitution as an example (α = 0.968), n is two, both atoms are at reactive

position (x = 2) and both of them are in competition for reaction (z = 2). Using

Eq. 25.22, an AKIE of 1.068 is obtained, which is close to the expected KIE for

cleavage of a C–Cl bond indicating that commitment to catalysis is small. While

the calculation of AKIE values using Eq. 25.22 provides considerable insight into

the causes of variations in isotope fractionation; it has some limitations. It is based

on the assumption that the molecule can be subdivided into reactive and non-reactive positions, which is not always the case. In some reactions, the stiffness of several bonds might vary causing isotope effects at multiple positions. Which bonds

change by how much is often not known unless methods of computational chemistry are applied to characterize the transition state.

In the context of research projects or practical applications, it can be of considerable interest to understand by which mechanism a compound is transformed.

For this purpose, observed isotope fractionation factors can be compared to those

from well-constrained laboratory studies or with calculated values. However, as

discussed above, the magnitude of isotope fractionation can vary somewhat for a

given mechanism, for example due to masking effects. Furthermore, under field

conditions, it is usually difficult to quantify isotope fractionation factor as concentrations are influenced by other processes than the transformation reaction.

Therefore, multi-element isotope approaches are generally preferred to identify

reaction mechanisms. As different mechanisms usually act on different bonds in

the molecules, the relative shift of the isotope composition among different elements varies, as illustrated in Fig. 25.3 for degradation of 1,2-DCA by two different mechanisms. Most often, such differences are illustrated in the form of

dual-element isotope plots and referred as a dual-element isotope approach.

Sometimes also the terms 2D plots and 2D isotope approach are used, which can

however create confusion, as in environmental studies 1D/2D/3D is generally used

to characterize the spatial dimension of a study or a modelling approach. A key

advantage of the method is that dual-element isotope slopes are generally not sensitive to masking because commitment to catalysis affects the isotopes of all elements to the same degree.



600



D. Hunkeler



25.6 Isotope Fractionation of Chlorinated Hydrocarbons

In this section, the effect of reactive processes on the isotope ratio of chlorinated

hydrocarbons is illustrated for different compound classes that commonly occur as

environmental contaminants.



25.6.1 Chlorinated Ethenes

Among chlorinated hydrocarbons, isotope fractionation has been studied most

extensively for chlorinated ethenes due to the prevalence of PCE and trichloroethene (TCE) as environmental contaminants. Isotope fractionation factors are

known for both biotic and abiotic transformation for parent compounds as well

as the most common intermediates. Biotic hydrogenolysis is particularly well

investigated as it commonly occurs in natural systems and is increasingly used as

a remediation method. The magnitude of carbon isotope fractionation (Table 25.3)

tends to increase from PCE to vinyl chloride (VC). In addition, for PCE and TCE

carbon isotope fractionation is more variable than for cis-1,2-dichloroethene (cis1,2-DCE) and VC. This likely reflects that PCE and TCE can be transformed by

a broader range of bacteria that might use different reaction mechanisms and/

or show various degree of “masking”. The enrichment factor for hydrogenolysis

of cis-1,2-DCE and VC correspond to AKIE that are close to the expected value

for cleavage of a C–Cl bond without masking. Less data are available for chlorine isotopes as analytical methods have only become available more recently. The

reported chlorine enrichment factors for hydrogenolysis (Table 25.3) are smaller

than for carbon, which is expected, given the smaller relative mass difference for

chlorine compared to carbon. Aerobic oxidation of VC and cis-1,2-DCE is associated with fairly consistent carbon isotope fractionation. The smaller isotope

enrichment factors compared to hydrogenolysis can be explained by more limited bond-modifications during formation of an epoxide intermediate compared to

bond cleavage during hydrogenolysis. Carbon and chlorine isotope fractionation

also occur during abiotic reductive dechlorination. The variable carbon isotope

enrichment factors probably originate from masking by rate limiting steps preceding reaction, whose rates may vary depending on the characteristics of metal surfaces. Finally, also oxidation of chlorinated ethenes by permanganate is associated

by substantial carbon isotope fractionation. Theoretical simulations have indicated

that the large carbon isotope fractionation for TCE originates from a concerted

reaction at both carbons by a 3 + 2 electrocyclic addition (Adamczyk et al. 2011).

For identifying transformation processes, dual-element isotope slopes are of particular interest (Fig. 25.5). Dual-element isotope slopes are fairly variable for PCE

degradation, likely again reflecting the diversity of bacteria and reaction mechanisms involved in their transformation. In contrast, the slopes seem to be more

consistent for TCE, cis-1,2-DCE and VC although the data set is still small. For

hydrogenolysis and abiotic reductive dechlorination of TCE, the slopes are fairly



25  Use of Compound-Specific Isotope Analysis (CSIA) …



601



Table 25.3  Carbon and chlorine isotope fractionation during biotic and abiotic transformation

of chlorinated ethenes by different mechanisms

Compound



Tetrachloroethene



Biotic

Oxidation

Reductive

dechlorina- (metabolic/

cometabolic)

tion

C



Cl



Trichloroethene



C



Cl



cis-1,2-Dichloroethene



C



Cl

Vinyl chloride



C



Cl



−3.6

(−0.4

to −19,

n = 31)

−2.0

(−0.9 to

−10.0,

n = 5)

−12.2

(−2.2 to

−18.9,

n = 31)

−4.3

(−2.7 to

−5.7, n = 7)

−18.8

(−12 to

−29.7,

n = 17)

−1.5

(n = 1)

−23.5

(−19.9

to −31.1,

n = 14)

−1.7

(n = 1)



Abiotic

Reductive

dechlorination

(Fe0, FeS, FeS2,

green rust)

−15.5

(−5.7 to −30.2,

n = 9)



Oxidation Oxidation

(permanga- (persulfate)

nate)

−17.0

(n = 1)



−25.1

−1.1 and

−16.7

−18.2, n = 2 (−7.5 to −33.4, (−21.4

to −26.8,

n = 17)

n = 3)

−2.6 (n = 1)

−8.4

(−0.4 to

−9.8, n = 6)



−21.1

−15.9

(−6.9 to −21.7, (n = 1)

n = 7)



−4.9

(n = 1)



−3.6

(n = 1)



−7.6

(n = 1)



−6.2

(n = 1)

−16.6

−6.3

(−6.9 to −19.4,

(−3.2 to

−8.2, n = 17) n = 5)

−0.3 (n = 1)



The median and range of the reported isotope enrichment factors are provided (Nijenhuis et al.

2005; Badin et al. 2014; Abe et al. 2009b; Audí-Miró et al. 2013; Barth et al. 2002; Chartrand

et al. 2005; Chu et al. 2004; Cichocka et al. 2007; Cretnik et al. 2013, 2014; Elsner et al. 2008;

Hunkeler et al. 1999, 2002, 2003; Lee et al. 2007; Numata et al. 2002; Poulson and Naraoka

2002; VanStone et al. 2004; Dong et al. 2009; Fletcher et al. 2011; Liang et al. 2007; Liu et al.

2014; Marchesi et al. 2012; Tiehm et al. 2008)



similar while they are more distinct for cis-1,2-DCE. Clearly, different slopes occur

for hydrogenolysis versus aerobic oxidation of cis-1,2-DCE and VC (Fig. 25.5).

The classical Rayleigh equation describes stable isotope patterns for the reactant and product of a one-step reaction. However, hydrogenolysis of chlorinated

ethenes proceeds step-wise with accumulation of less-chlorinated intermediates.

Isotope patterns of intermediates are often influenced by formation and degradation of the compound. Isotope patterns for sequential hydrogenolysis of



602



D. Hunkeler



Fig. 25.5  Dual-element isotope slopes Λ (C/Cl) for various transformation processes of chlorinated ethenes (Abe et al. 2009a; Audí-Miró et al. 2013; Cretnik et al. 2014, 2013; Kuder and Philp

2013; Renpenning et al. 2014)



chlorinated ethenes are distinctly different depending on the element as for carbon, all atoms remain conserved, for chlorine, atoms are cleaved, while for hydrogen, atoms are added. For carbon, each intermediate is depleted in light isotopes

initially compared to the precursor (Hunkeler et al. 1999), but depending on the

relative rate of degradation and the enrichment factors can get heavier than the

precursor as degradation advances (Van Breukelen et al. 2005). If an intermediate is not degraded further, its carbon isotope ratio approaches that of the initial

compound for mass balance reasons. Only if the intermediate gets degraded further can it become “heavier” than the initial compound. For chlorine isotopes, the

isotope pattern is more complex as chlorine atoms are released (Hunkeler et al.

2009). As the released chlorine atoms are depleted in heavy isotopes, the intermediate becomes enriched in heavy isotopes relative to the starting point, even if

the intermediate does not degrade further. In addition, a secondary isotope effect

during hydrogenolysis can further shift the isotope ratio of the remaining chlorine

atoms in the intermediate (Cretnik et al. 2014). As for hydrogen, the incorporated

atoms are fairly depleted in heavy isotope. As a result, each intermediate has more

depleted isotope signature compared to the precursor (Kuder et al. 2013).



25  Use of Compound-Specific Isotope Analysis (CSIA) …



603



25.6.2 Chlorinated Ethanes

Among the chlorinated ethanes, 1,2-DCA, 1,1,1-trichloroethane (1,1,1-TCA)

and 1,1,2,2-tetrachloroethane (1,1,2,2-PCA) occur frequently as environmental

contaminants. Under oxic conditions, 1,2-DCA can be biodegraded by two different mechanisms, via an SN2 nucleophilic substation or C–H bond oxidation.

As already discussed above, these mechanisms are associated with distinctly different isotope fractionation factors and dual-element isotope slopes (Fig. 25.3,

Table 25.4). Under anoxic conditions, biodegradation proceeds via dichloroelimination to ethene, a process that is associated with strong carbon isotope fractionation originating from cleavage of the two carbon–chlorine bonds (Table 25.4).

Alternatively, 1,2-DCA might biodegrade by hydrogenolysis to chloroethane or

undergo abiotic dehydrohalogenation to vinyl chloride. However, for these pathways, isotope fractionation has not been characterized yet.

1,1,1-TCA can be degraded abiotically and biotically. For abiotic degradation,

distinctly different dual-element isotope slopes were observed for hydrolysis/deh

ydrohalogenation, reductive dechlorination with Fe(0) and oxidation by heat-activated persulfate (Table 25.4). It is interesting to note that the former two mechanisms are associated with significantly different slopes although both of them

involve cleavage of a C–Cl bond in the initial transformation step. Secondary isotope effects might contribute to isotope fractionation to a different degree.



25.6.3 Hexachlorohexane

Hexachlorohexane (HCH) degradation is a good example to illustrate how the molecule size (“dilution”) and rate-limiting steps (“masking”) influence the magnitude

of isotope fractionation. Generally, carbon isotope enrichment factors (−0.7 to

−7.6 ‰) tend to be smaller than for chlorinated ethanes (Table 25.4). This can be

explained by the larger number of carbon atoms in the molecules which increases

the probability that the heavy isotope is located in a non-reacting position. Two

reaction mechanisms, dehydrochlorination and dichloroelimination occur during

abiotic as well as biotic degradation of the compound. In both cases, the enrichment factors are larger during abiotic than biotic degradation. Rate-limiting transport to the enzyme or enzyme-substrate binding might partly mask the KIE.



25.6.4 Chlorinated Benzenes

Chlorinated benzenes have been used as degreasers, solvents, pesticides and for

chemical synthesis. Under aerobic conditions, chlorinated benzenes are generally

degraded via dihydroxylation by a ring dioxygenase (Van der Meer et al. 1991;

Field and Sierra-Alvarez 2008). During this process, carbon isotope fractionation is



−3.0 to −3.8

−32.1

−1.6

−7.8 to −13.6

−4.0



Aerobic oxidation C–H cleavage



Dichloroelimination

Hydrolysis and dehydrochlorination

Abiotic reduction by Fe(0)

Oxidation by heat-activated persulfate



γ-Hexachlorohexane



α-Hexachlorohexane



Anaerobic biodegradation via

dichloroelimination

Anaerobic biodegradation



Aerobic biodegradation via

dehydrochlorination



Abiotic reduction by Fe(0) via

dichloroelimination



Abiotic reduction by Fe(0)

Alkaline hydrolysis via

dehydrochlorination



−3.4 to −3.9







Badea et al. (2011)







Badea et al. (2009)



Bashir et al. (2013)











Zhang et al. (2014)







εCl (‰)

Λ (C/Cl) Refs.

−4.2 to −4.4 7.6–7.7 Hirschorn et al. (2004), Hunkeler and

Aravena (2000)

−3.8

0.78

Hirschorn et al. (2004), Palau et al.

(2014a)





Hunkeler et al. (2002)

−4.7

0.33

Palau et al. (2014b)

−5.2

1.5

Elsner et al. (2007), Palau et al. (2014b)

No significant ∞

Palau et al. (2014b)

fractionation





Marchesi et al. (2013)





Sherwood Lollar et al. (2010)





Elsner et al. (2007), Hofstetter et al.

(2007), Neumann et al. (2009)





Elsner et al. (2007)





Zhang et al. (2014)



−19.3

−7.6 bulk

−1.7 (+) enantiomer

−2.1 (−) enantiomer



−4.9 bulk

−5.1 (+) enantiomer

−4.8 (−) enantiomer



−1.0 to −1.6 bulk

−2.4 to −2.5 (+)

enantiomer

−0.7 to −1.0 (−)

enantiomer

−3.7 bulk





Oxidation by base-activated persulfate −7.0

1.8

Biotic hydrogenolysis

−20.8 to −27.1

1,1,2,2-Tetrachloroethane Dehydrochlorination



1,1,1-Trichloroethane



1,2-Dichloroethane



εC (‰)

−28.7 to −33.4



Mechanism

Aerobic oxidation SN2



Table 25.4  Isotope enrichment factors and dual-element isotope slopes for environmentally relevant degradation processes of common chlorinated ethanes and hexachlorohexane



604

D. Hunkeler



25  Use of Compound-Specific Isotope Analysis (CSIA) …



605



Table 25.5  Isotope enrichment factors for biodegradation of chlorinated benzenes

Compound

Monochlorobenzene

Monochlorobenzene

1,2-Dichlorobenzene

1,3-Dichlorobenzene

1,4-Dichlorobenzene

1,2,3-Trichlorobenzene



Mechanism/conditions

Aerobic—pure culture

Anaerobic

Anaerobic

Anaerobic

Anaerobic

Anaerobic



εC (‰)

−0.1 to −0.4

−5.0

−0.8

−5.4

−6.3

−3.5



1,2,4-Trichlorobenzene



Anaerobic



−3.0 to −3.2



Refs.

Kaschl et al. (2005)

Liang et al. (2011)

Liang et al. (2014)

Liang et al. (2014)

Liang et al. (2014)

Griebler et al. (2004),

Liang et al. (2011)

Griebler et al. (2004),

Liang et al. (2011)



absent or very small, as shown for monochlorobenzene and 1,2,3-trichlorobenzene

(Table 25.5). Similar observations have also been made for the degradation of other

aromatic compounds, such as toluene, by ring dioxygenases (Vogt et al. 2008).

The small or absent carbon isotope fractionation can be explained by the absence

of complete bond cleavage during the dihydroxylation process. Under anaerobic

conditions, mono- and polychlorinated benzenes can be degraded by hydrogenolysis similarly as chlorinated ethenes (Adrian et al. 2000; Field and Sierra-Alvarez

2008). Significant carbon isotope fractionation occurs with enrichment factors

reaching from −3.0 to −6.3 ‰ (Table 25.5), except for 1,2-dichlorobenzene,

which shows less isotope fractionation (−0.8 %). The fairly large isotope fractionation can be explained by C–Cl bond cleavage in the rate limiting step. The magnitude of carbon isotope fractionation is, however, smaller than for most chlorinated

ethenes. This can again be explained by the larger number of carbon atoms in nonreactive positions in chlorinated benzenes compared to chlorinated ethenes.



25.6.5 Chlorinated Methanes

Chlorinated methanes have been widely used as solvents and reagents in organic

synthesis. In addition, some chlorinated methanes can also be naturally produced or

occur as a by-product during chlorination of drinking water. Chloromethane is the

most abundant volatile chlorinated hydrocarbon in the atmosphere, most of which

originates from terrestrial vegetation (Keppler et al. 2005). Most of the natural chloroform originates from biogenic sources such as plants, insects and probably fungi,

with comparable contributions by terrestrial and oceanic environments (Laturnus

et al. 2002). In groundwater below coniferous forests, chloroform can occur naturally at concentrations up to 10 µg/L (Hunkeler et al. 2012). Biodegradation and

abiotic transformation of chlorinated methanes tends to be associated with a large

carbon isotope fractionation (Table 25.6). As only one carbon is present, no reduction of the KIE by non-reactive positions occurs. For chloromethane, a C–Cl bond

is broken by a methyl transfer reaction in the initial step (Vannelli et al. 1998). This



D. Hunkeler



606

Table 25.6  Isotope enrichment factors for degradation of chlorinated methanes

Compound

Chloromethane



Dichloromethane



Dichloromethane



Chloroform



Carbon

tetrachloride



Mechanism/

conditions

Biotic

Aerobic by methylotrophic bacteria

Biotic

Aerobic by methylotrophic bacteria



εC (‰)



−42.4 to

−66.3



−3.8







Biotic

Denitrifying by

methylotrophic

bacteria

Biotic

Dehalobactercontaining enrichment culture

Abiotic reductive

dechlorination by

Fe(II) in smectites

Abiotic reductive

dehalogenation by

iron (hydr)oxides or

siderite

Abiotic reductive

dehalogenation by

mackinawite

Alkaline hydrolysis



−45.8 to

−61.0











−27.5











−38 to −50



εCl

(‰)





εH (‰)



Refs.



−27 to −29



Nadalig et al.

(2013), Miller

et al. (2001)

Heraty et al.

(1999),

Nikolausz et al.

(2006)

Nikolausz et al.

(2006)



Chan et al.

(2012)



−10.9 to

−13.3



Neumann et al.

(2009)



−25.6 to −32



Zwank et al.

(2005)



−15.9



Zwank et al.

(2005)



−49 to −56



Torrento et al.

(2014)



also explains why only small hydrogen isotope fractionation occurs likely due to

a secondary hydrogen isotope effect. For dichloromethane degradation by methylotrophic bacteria, the initial transformation likely occurs via a biologically mediated nucleophilic substitution which also involves cleavage of a C–Cl bond. This

proposed mechanism is consistent with the relatively large chlorine isotope enrichment factor of −3.8 ‰ (Heraty et al. 1999). For carbon tetrachloride, large but

variable carbon isotope fractionation was observed for reductive dechlorination by

solid phases containing reduced iron or sulfur. The variation in isotope fractionation

might be due to “masking” by rate-limiting transport to reactive solid phase.



25.7 Application of Isotope Methods in Field Studies

Stable isotope methods can be applied to address different questions related to

the origin and fate of chlorinated hydrocarbons in the environment (Table 25.7).

Applications can broadly be divided in those that aim at identifying different



Chloroform from natural production in forest soils versus anthropogenic chloroform (Hunkeler et al. 2012)



Comparing isotope signatures

to known ranges for natural and

anthropogenic compounds

Determination of dual-element

isotope slopes for field samples

and comparison with laboratory

value

Evaluation of shifts in isotope

ratios

Evaluation of isotope patterns of

reactants and products



Field dual carbon-chlorine isotope slopes suggested that

cis-1,2-dichloroethene was subject to abiotic biodegradation

(Hunkeler et al. 2011a)



Example

Dual carbon–chloride isotope analysis indicated presence of multiple trichloroethene sources in bedrock aquifer (Lojkasek-Lima

et al. 2012)



Approach

Comparison of multi-element

isotope signatures in source area

and downgradient plumes



Tracking progress of a process

qualitatively

Linking reactants to products of

degradation for contaminant mixtures where multiple pathways can

lead to the same product



Demonstrating natural attenuation of chlorinated ethenes based

on enrichment of heavy carbon isotopes (Hunkeler et al. 1999).

Carbon isotope analysis indicated that TCE originated partly

from dehydrochlorination of 1,1,2,2- tetrachloroethane. Vinyl

chloride and ethene were identified as products of dehydrochlorination of 1,1,2-trichloroethane and dichloroelimination

of 1,2-dichloroethane rather than reductive dechlorination of

chlorinated ethenes (Hunkeler et al. 2005)

Estimation of the amount of chlorinated hydrocarbon transformaSimplified estimation of degree of Application of the Rayleigh

equation based on known isotope tion by natural attenuation (Vieth et al. 2003)

contaminant transformation

enrichment factors from laboratory studies

Estimation of rate of cis-1,2-dichloroethene biodegradation at

Estimation of first order transfor- Combining Rayleigh equation

with first order rate expression to site with bioaugmentation (Morrill et al. 2005)

mation rate

estimate rate constants relying on

laboratory enrichment factors

Isotope data of tetrachloroethene and trichloroethene made it

Quantification of reaction progress Reactive transport model that

possible to distinguish between two reactive transport models

and prediction of future evolution incorporates isotope data

that reproduced product patterns well but gave different predictions for future trends (Atteia et al. 2008)



Tracking reactive Identification by which process a

contaminant is degraded at a site

processes



Objective

Tracking sources Evaluating if different sources of

of contaminants the same compound are present

at a site and linking downgradient

contamination to their source

Distinguishing natural and anthropogenic sources of a compound



Table 25.7  Common applications of isotope methods in field studies



25  Use of Compound-Specific Isotope Analysis (CSIA) …

607



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