Science of the Total Environment 514 (2015) 459–466

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Enantioselective stable isotope analysis (ESIA) — A new concept to evaluate the environmental fate of chiral organic contaminants Silviu-Laurentiu Badea a,⁎, Andrei-Florin Danet b a b

Department of Chemistry, Umeå University, SE-901 87 Umeå, Sweden Department of Analytical Chemistry, University of Bucharest, Faculty of Chemistry, 90-92 Panduri Str., Bucharest 050657, Romania

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• ESIA is an innovative technique to assess the environmental fate of chiral pollutants • Overcoming the analytical limitations of ESIA is challenging • Development of ESIA methods for new chiral emerging contaminants is needed

a r t i c l e

i n f o

Article history: Received 24 November 2014 Received in revised form 26 January 2015 Accepted 26 January 2015 Available online xxxx Editor: E. Capri Keywords: ESIA CSIA Enantioselective analysis CSP

a b s t r a c t Since 2011, the enantiospecific stable carbon isotope analysis (ESIA) has emerged as an innovative technique to assess the environmental fate of chiral emerging compounds by combining in one experimental technique both compound specific isotope analysis (CSIA) and enantioselective analysis. To date, the ESIA was applied for four classes of compounds: α-hexachlorocyclohexane (α-HCH), polar herbicides (phenoxy acids), synthetic polycyclic musk galaxolide (HHCB), and phenoxyalkanoic methyl herbicides. From an analytical point of view there are factors that are hindering the application of ESIA methods for the field samples: (i.e. amounts of target analyte, matrix effects, GC resolution) and overcoming these factors is challenging. While ESIA was shown as a mature technique for the first three abovementioned class of compounds, no isotope analysis of individual enantiomers could be performed for phenoxyalkanoic methyl herbicides. With respect to field studies, one study showed that ESIA might be a promising tool to distinguish between biotic and abiotic transformation pathways of chiral organic contaminants and even to differentiate between their aerobic and anaerobic biotransformation pathways. The development of ESIA methods for new chiral emerging contaminants in combination with development of multi-element isotope analysis will contribute to a better characterization of transformation pathways of chiral organic contaminants. © 2015 Elsevier B.V. All rights reserved.

Contents 1. 2.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Development of enantioselective carbon isotope analysis (ESIA) analytical methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.1. Challenges and limitations of ESIA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

⁎ Corresponding author. E-mail address: [email protected] (S.-L. Badea).

http://dx.doi.org/10.1016/j.scitotenv.2015.01.082 0048-9697/© 2015 Elsevier B.V. All rights reserved.

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ESIA as tool to characterize biodegradation of organic contaminants 3.1. α-Hexachlorocyclohexane . . . . . . . . . . . . . . . 3.2. Phenoxy acids . . . . . . . . . . . . . . . . . . . . . 3.3. Galaxolide (HHCB) . . . . . . . . . . . . . . . . . . . 3.4. Phenoxyalkanoic methyl herbicides . . . . . . . . . . . 4. ESIA as a differential tool between abiotic and biotic transformation of organic contaminants . . . . . . . . . . . . 5. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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1. Introduction Currently, the stable isotope methodology is a key technology improving the understanding of biogeochemistry of organic chemicals since it has been increasingly considered for characterizing in situ biodegradation processes and also used to assess the role of microbial communities in these processes (Meckenstock et al., 2004). Compound specific stable isotope analysis (CSIA) is taking advantage of the preferential transformation of lighter isotopomers during a degradation reaction (due to the preferentially cleavage of the bonds 12C\Cl comparing with the bonds 13 C\Cl), thus leading to an enrichment of heavier isotopes in the residual phase in the course of biodegradation (for an overview of the method see Elsner et al., 2005; Meckenstock et al., 2004). In this respect, the Rayleigh concept has been widely used in environmental studies (see the review paper Meckenstock et al., 2004) to relate the change of bulk isotope ratios of chemicals to the extent of the target compound degradation. The subject of compound specific isotope analysis (CSIA) for environmental investigations has been an important step forward in contemporary environmental science. Concepts and applications are currently available for carbon and hydrogen based CSIA of the more simple organic contaminants such as BTEX (Fischer et al., 2008), chlorinated ethenes (Cichocka et al., 2007; Lollar et al., 2001; Marco-Urrea et al., 2011; Nijenhuis et al., 2005) and MTBE (Rosell et al., 2007), and are already integrated into common monitoring strategies for polluted sites. On the contrary, applications and concepts for larger molecules comprising compound classes such as brominated flame retardants, pesticides, persistent organic chemicals (POPs) and emerging chemicals are still in its early stages. Currently, only scattered data obtained in the CSIA of POPs are available, e.g. for polychlorinated biphenyls (PCBs) (Drenzek et al., 2001) and (Horii et al., 2005), polychlorinated naphthalenes (PCNs) (Horii et al., 2005), DDT (Reddy et al., 2002), toxaphene (Vetter et al., 2005), hexachlorocyclohexanes (Badea et al., 2009), polychlorinated dibenzo-p-dioxins (Horii et al., 2008) and (Liu et al., 2010), and polybrominated diphenyl ethers (PBDEs) (Vetter et al., 2008). Thus, methodological developments are urgently needed to exploit the potential of CSIA in environmental research. Many organic pollutants (i.e. o,p′-DDT, α-HCH, chiral PCBs, chiral polybrominated biphenyls (PBBs), etc.) have a chiral carbon skeleton and they are commonly released into the environment as racemate mixtures and can undergo enantiomer specific decomposition during microbial or chemical reactions in the environment. Generally, it is difficult to predict which enantiomer may be enriched since accumulation of an enantiomer is dependent on the environmental system. It is believed that enantioselective degradation of a contaminant implies that the enzymes involved in the conversion of such compounds are able to differentiate between the enantiomers (Muller and Kohler, 2004). This means that one enantiomer is degraded faster than the other (Harrison et al., 1998; Zipper et al., 1998a, 1999, 1998b), either due to preferential microbial uptake or preferential enzyme activity. Very recently, Qiu et al. (2014) suggested that enantiomer fractionation might occur through a different mechanism comparing with isotope

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fractionation. Naturally, isotope fractionation is primarily expected to appear if chemical bonds are changed in the enzymatic transformation (Elsner, 2010) but not during cell uptake. In contrast, as mentioned earlier, the enantiomer fractionation can occur either through transporter-driven cell uptake or through enzymatic reactions. Thus in addition to CSIA, enantiospecific analysis of chiral organic pollutants opens opportunities for characterizing biodegradation processes since enantiomers of chiral compounds often exhibit differences in biodegradation rates as most biochemical processes in nature are stereospecific (Padma et al., 2003a). Further enantiomeric patterns of organic pollutants (described as enantiomeric fractions (EFs)) may be used as fingerprint for their origin enabling investigation of processes, long-range transport and deposition. Nevertheless, both CSIA and enantioselective analysis have limitations when applied to environmental research (see Table 1). For CSIA, distinct isotopic compositions of pollution sources are required, and the differences between sources must be greater than the measurement error (usually ± 0.5‰). In addition, the observed difference in the values of isotope compositions must exceed the spatial and temporal variability introduced by different sources of pollution at the site, by the mixing of ground water flow lines, and by sorption and volatilization. Moreover, interpretation becomes complex when there are many possible sources (Elsner et al., 2012; Oulhote et al., 2011). Thus, to ensure a reliable interpretation, the difference in the δ13C values must be at least 1‰. Nevertheless, in many field studies a difference of at least 2‰ is recommended. In enantiomeric analysis, as reported in previous studies (Badea et al., 2011; Padma et al., 2003b) substantial biodegradation may occur without significant deviations of the enantiomeric fractions (EFs). Thus, EF values are not always a reliable tool for assessing the biodegradation of certain contaminants. In recent years, the combination of enantiospecific analysis and CSIA has become a promising new approach that can provide insight into enantioselective fates and source apportionment of environmental organic contaminants. It can overcome to some extent the limitation of using only a single technique. Badea et al., 2011 developed an enantiospecific stable carbon isotope analysis (ESIA) method for α-HCH enantiomers, and three dense, nonaqueous-phase liquids obtained from an HCH-contaminated field site were analyzed to test the applicability of the method. They found that the isotopic compositions of the α-HCH enantiomers showed a range of enantiomeric and isotope patterns, suggesting that enantiomeric and isotope fractionation can serve as an indicator for biodegradation and source characterization of α-HCH in the environment. Very recently, individual isotope enrichment factors of α-HCH enantiomers during aerobic biodegradation of α-HCH have been reported by Bashir et al. (2013). This ESIA was applied afterwards by Milosevic et al. (2013) and Maier et al. (2013) for polar herbicides (phenoxy acids), Wang et al. (2013) for galaxolide (HHCB), a synthetic polycyclic musk and Jammer et al. (2014) for phenoxyalkanoic methyl herbicides. The approach gives perspectives for assessing the environmental fate of any chiral xenobiotic compound. The objective of this paper is an overview of enantioselective isotope analysis (ESIA) methods currently developed for organic contaminants, both from analytical and environmental points of view.

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Table 1 Comparison between chiral analysis, CSIA and ESIA. Characteristics

Chiral analysis

CSIA

ESIA

Molecular weight of the target compounds Analysis sensitivity

Analysis can be used for large compounds Requires pg of analyte

Restricted to rather small compounds ( C12)

Restricted to rather small compounds ( C12) Requires ng of analyte

Chromatographic resolution

Requires baseline separation (R ≥ 1.5)

Analytical error

±0.002 EF units

Requires ng of analyte (about 1000-fold higher that for GC chiral analysis). Limit of detections in ng of compound to be delivered on GC column for the selected target compounds varied between 14.35 ng for galaxolide (HHCB) and 48.47 ng for α-HCH. Requires baseline separation (R ≥ 1.5)

±0.5‰

2. Development of enantioselective carbon isotope analysis (ESIA) analytical methods Enantioselective analysis of chiral organic compounds is preferably performed with high-resolution gas chromatography (HRGC) using modified cyclodextrin (CD) as a chiral stationary phase (CSP) (Eljarrat et al., 2008; Schurig, 1994). The hydroxy groups at positions 2-, 3- (at the wide end) and 6- (at the narrow end) can be derivatized (usually permethylated) to form modified cyclodextrins with different selectivity and better thermal stability than pure CD. Nevertheless, the enantiomer separations of cyclodextrins is given by both their architecture and chemistry since the chemical interactions that lead to chiral separations can occur on both the exterior and interior surfaces of the cyclodextrin toroid. The most important consideration for retention and chiral recognition is proper fit of the analyte into the cyclodextrin cavity. This fit is a function of both molecular size and shape of the analyte, relative to the cyclodextrin cavity. For example, the large analytes (i.e. α-HCH) are usually better separated on permethylated γ-cyclodextrin. There are two basic mechanisms involved in chiral separations on cyclodextrins; those that occur on the inside cavity surface (inclusion complexing, Fig. 1A), and those that occur on the outside surface (surface interactions, Fig. 1 B) of the cyclodextrin toroid. Thus, the development of enantioselective stable isotope analysis (ESIA) methods involved gas chromatography–combustion–isotope ratio mass spectrometry (GC–C– IRMS) analysis performed with chiral GC column based on modified cyclodextrins. The choice of chiral GC column depends on the target chiral contaminants to be investigated. For example, Maier et al. (2013) performed the ESIA of phenoxy acids on a permethylated β-cyclodextrin GC column (β-6TBDM, column length × I.D. 50 m × 0.25 mm, df = 0.25 μm, Macherey & Nagel, Düren, Germany), while Badea et al. (2011) used a permethylated γ-cyclodextrin GC column (γ-DEX™ 120, column length × I.D. 30 m × 0.25 mm, df = 0.25 μm, Supelco, containing 20% permethylated γ-cyclodextrin) for ESIA of α-HCH, since it was shown (Hardt et al., 1994) that permethylated γ-cyclodextrin GC columns have

Requires a very good chromatographic resolution (R ≥ 3.0) ±0.5–0.75‰

a higher enantioselectivity for separation of α-HCH enantiomers compared to other chiral columns based on permethylated α-cyclodextrin and β-cyclodextrin (Buser and Muller, 1995; Muller et al., 1992; Suar et al., 2005). Another factor to be considered in ESIA are the parameters of the GC system. The GC oven temperature plays a critical role for the enantioselective separation of compounds on modified cyclodextrin CD. Generally, enantiomers are better separated and thus higher separation factors (α) are obtained at low temperatures (160 °C or lower) (Badea et al., 2011) and high average velocities (about 50 cm/s) since the chiral separation is governed by entropy. Comparing with the normal GC–C–IRMS analysis, in ESIA, the analytical signal of chiral contaminant is spitted in two distinguish signals proportionally with the amount of each enantiomer injected in the GC column and passed through the combustion furnace (see Fig. 2). Therefore, for calculations of the isotopic composition of a bulk chiral contaminant X from the isotope ratios of his individual enantiomers, the following equation can be used as demonstrated by Badea et al. (2011): 13

13

13

δ C X ½‰ ¼ EFð−Þ  δ C ð−ÞX ½‰ þ EFðþÞ  δ C ðþÞX ½‰

ð1Þ

where EF(−) and EF(+) are the enantiomer-selective fractionation of X enantiomers. As usually, the carbon isotope ratios were reported in δ notation in parts per thousand (‰) relative to the international carbon isotope standard Vienna Pee Dee Belemnite (V-PDB) (Coplen et al., 2006), according to the following equation: 13

δ C X ½‰ ¼

  Rsample −Rs tandard  1000 Rs tandard

ð2Þ

where Rsample is the carbon isotope ratio (13C/12C) of the sample and Rstandard is carbon isotope ratio (13C/12C) of the VPDB-standard. In Eq. (1), the enantiomeric fraction (EF) is used to describe the relationship between enantiomers during biodegradation (de Geus et al., 2000; Harner et al., 2000; Wiberg et al., 2001a, 2001b) and is usually measured by chiral GC–MS analysis due to higher precision in measuring peak areas of a GC–MS system comparing with GC–C–IRMS. The EF(+) is defined as A+ / (A+ + A−), where A+ and A− correspond to the peak areas of the (+) and (−) enantiomers (Harner et al., 2000). Racemic compounds has an EF(+) equal to 0.5. An EF(+) N 0.5 indicates preferential degradation of the (−) enantiomer, and an EF(+) b 0.5 indicates preferential degradation of (+) enantiomer. The enantiomeric fraction can be also determined relating to the (−) enantiomer; EF(−) is then defined as A− / (A+ + A−). The evaluation of EF(−) leads to the opposite interpretation as given for EF(+). 2.1. Challenges and limitations of ESIA

Fig. 1. Interactions of cyclodextrin with the analyte: Inclusion complexing mechanism of an analyte into the cyclodextrin cavity (A). Surface interaction mechanism of an analyte with the cyclodextrin (B).

There are three main factors that are hindering the wide application of ESIA methods for the field and even artificial matrix samples: I.

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Sample He

(-) (+)

CO2 CO2 Combustion Furnace

GC chiral separation of enantiomers on permethylated α, β or γ-cyclodextrin

Magnet

CO2

m/z 44 (12C16O2)

Ion source

m/z 45 m/z 46 13 16 (12C18O16O) ( C O2) Standard

Valve

Detector

CO2 of known isotope composition Fig. 2. ESIA performed on a GC–C–IRMS for carbon isotope analysis.

Concentrations of target analytes required; II. Matrix effects; and III. Resolution of the GC system (see Table 1). I. In environmental matrices, most organic contaminants and their metabolites are present only in trace concentrations (in the ng/mL or ng/g range). Taking into account that it is necessary to inject approximately 1 nmol carbon on column, the limit of detections of the GC–IRMS system in ng of compound to be delivered on GC column for the selected target compounds varied between 14.35 ng for galaxolide (HHCB) and 48.47 ng for α-HCH (see Table 1). However, due to the experimental conditions of the isotope analysis method (GC parameters, IRMS tunning, etc.), a more than 1000-fold higher concentration (i.e. 10–100 μg/mL) is required for CSIA or ESIA (la Farre et al., 2008). Concentration alone can provide larger amounts of analyte but cannot resolve matrix interferences. Therefore, in many cases, a preconcentration procedure is needed for an accurate determination of isotope composition of organic contaminants and their metabolites in environmental matrices. II. Most chiral gas chromatography (GC) columns have relatively low maximum operating temperatures (for example, b 230 °C) compared with common GC columns, and many matrix interfering substances are, therefore, not eluted from the chiral columns (Table 1). These matrix effects and column bleed might reduce sensitivity and also add to the number of combustion products when conducting gas chromatography/isotope ratio mass spectrometry (GC/IRMS) analysis (Elsner et al., 2012; Meyer et al., 2008). Thus, an efficient preparative procedure (clean-up) that will reduce the matrix effects is needed for field samples (wastewater, soil, biota, etc.) (Amaral et al., 2010; Berg et al., 2007; Meyer et al., 2008; Penning and Elsner, 2007). III. Chiral GC–C–IRMS analysis requires the avoiding of any interferences with the target analyte (baseline separation needed). However, there are many GC–C–IRMS applications where closely related compounds simply cannot be resolved resulting in peaks overlap. Actually, CO2 and N2 disperse more freely within the carrier gas stream than their parent organic compounds which can result in overlapping CO2 peaks even for baseline resolved GC peaks. Furthermore, the chromatographic separation in terms of optimum peak-shape and baseline separation is likely to be impaired during combustion and the subsequent passage through the interface. Also, the changes in tubing diameter and frequent use of unions to connect the various parts of tubing might lead to peak broadening and peak distortion and even to partial peak overlap. All of the above mentioned factors can affect the accuracy and precision of isotope ratio measurements (Meier-Augenstein et al., 1996). Since many compounds are presented in the field samples, the optimisation of GC parameters used in ESIA analysis of the

field samples is technically challenging. For example, using the same GC temperature program and the same flow rate (1.5 mL/min), Badea et al., 2011 observed on the GC–MS system a difference between the retention times of the α-HCH enantiomers about 1 min, while on the GC–C–IRMS system the difference between the isotopic signatures of the α-HCH enantiomers was only about 20 s, still allowing a very good baseline separation. Thus very slow GC temperature ramp or a prolonged isothermal phase must be used in ESIA.

3. ESIA as tool to characterize biodegradation of organic contaminants 3.1. α-Hexachlorocyclohexane Until now, two studies were published on quantification of α-HCH biodegradation using ESIA. One of the study (Badea et al., 2011) demonstrated the applicability of the innovative enantioselective stable isotope analysis (ESIA) concept for the anaerobic biotransformation with Clostridium pasteurianum. Anaerobic lab experiments performed by Badea et al. (2011) showed almost non-enantioselective biodegradation of α-HCH by C. pasteurianum indicating the potential of the distinction for aerobic and anaerobic α-HCH biodegradation using EF (+) or EF (−) values. Furthermore, ESIA was applied for the determination of α-HCH enantiomers in 3 field DNAPL (dense non-aqueous phase liquid) samples and by comparing these field isotope values with a racemic standard, the isotopic composition of a bulk α-HCH in one nonracemic sample (D21) was calculated from the isotope ratios of these individual enantiomers (− 25.6 ± 0.3‰ for (−) α-HCH and − 26.5 ± 0.1‰ for (+) α-HCH respectively) using the abovementioned Eq. (1), thus showing the accuracy of ESIA for real environmental samples (Fig. 3). In a subsequent study, Bashir et al. (2013) performed extensively aerobic degradation experiments of α-HCH by two Sphingobium spp. strains, the Sphingobium indicum strain B90A and the Sphingobium japonicum strain UT26, and determined the carbon stable isotope fractionation during aerobic degradation of the enantiomers compared to the isotope fractionation of bulk α-HCH. Furthermore, for the first time, individual isotope enrichment factors of α-HCH enantiomers during aerobic biodegradation of α-HCH were reported in this study. For both strains, the S. indicum strain B90A and the S. japonicum strain UT26, higher isotope enrichment factors for (+) α-enantiomer (− 2.4 ± 0.8‰ and − 2.5 ± 0.6‰, respectively) as compared to the (−) α-enantiomer (− 1.0 ± 0.6‰ and − 0.7 ± 0.2‰, respectively) were obtained. In concordance with the abovementioned Eq. (1), the

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Fig. 3. Adapted from Badea et al., 2011, Rapid Communications in Mass Spectrometry. The isotope ratios of (−) α-HCH (●) and (+) α-HCH (○) vs. EFs for dense non-aqueous phase liquid (DNAPL) samples with the code name D4, D21 and 049 compared to a racemic reference compound of α-HCH. D4 is racemic (marked with green), while D21 (red) and 049 (blue) are non-racemic. (For interpretation of the references to color in this figure legend, the reader is referred to the web version of this article.)

average of the carbon isotope enrichment factors for (+)- and (−)-αHCH enantiomers was within the statistical uncertainty of the εc value of bulk α-HCH. In the abovementioned study, the distinction in the carbon isotope fractionation might be caused by different extents of rate limitation preceding the isotope-sensitive step of the biodegradation of the enantiomers. Eventually, Bashir et al. (2013) suggested that a slower biodegradation induced a slower enzymatic turnover of (+) αenantiomer and this might cause a shift in the rate limitation from non-isotope fractionating, preceding steps toward the isotopesensitive step of the enzymatic bond cleavage leading to a more pronounced carbon isotope fractionation of (+) α-enantiomer. This explanation is supported by higher observed isotope fractionation for the (+) compared to the (−) α-enantiomer. Recently, Gasser et al. (2012) proposed to quantitatively describe the fractionation of enantiomers induced by biotransformation of contaminants using the Rayleigh equation, similarly to the quantification of biodegradation using isotope fractionation. Therefore, using this new concept, Bashir et al. (2013) calculated additionally the enantiomeric enrichment factors (εe) for α-HCH enantiomers. Bashir et al. (2013) concluded that the combination of enantiomer and carbon isotope fractionation, in a two-dimensional approach, using values of enantiomeric fraction for both enantiomers (EF(+) and EF(−)) in correlation with the stable carbon isotope discrimination (i.e. variation of isotope composition during field degradation) of the αHCH enantiomers should allow to trace degradation pathways of αHCH and in the environment (to distinguish between their biotic and abiotic transformation pathways in sediments and even to differentiate between their aerobic and anaerobic biotransformation pathways). This approach was applied a further study by Zhang et al. (2014). 3.2. Phenoxy acids Milosevic et al. (2013) published a first article on ESIA applied to the enantiomers of phenoxy acids to assess their fate in a heterogeneous geologic setting at an old landfill (Risby site). For the first time, the environmental fate of phenoxy acids in a contaminated landfill was assessed by a combination of four different methods was used: analysis of (i) parent and daughter compound concentrations, (ii) enantiomeric fractionation (expressed as enantiomeric excess (EE) (Patterson and Schnell, 2014; Shi et al., 2006)) (iii) compound-specific isotope analysis and (iv) enantioselective stable isotope analysis.

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The combined values of concentrations, enantiomeric excess and isotope ratios of dichlorprop and its putative metabolite 4-CPP (2-(4chlorophenoxy) propionic acid) confirmed their dechlorination in the hotspot and provided evidence for further degradation of 4-CPP downgradient of the hotspot. More specifically, the enantiomeric excess of 49% in 4-CPP and the difference between isotope composition of 4CPP enantiomers (up to 3‰) indicates (S)-4-CPP-preferential degradation in the hotspot. The study concluded that the combined enantiomer and isotope analysis in the hotspot indicated non-enantiospecific transformation of dichlorprop and confirmed preferential degradation of (S)4-CPP enantiomer under anaerobic conditions. However, the ESIA of (S)- to (R)-enantiomers of 4-CPP was not fully optimized (tough fairly good) and due to the uncertainty of isotope measurement (standard deviation ± 0.75‰) further studies were needed to assess the in situ transformation of phenoxy acids in the abovementioned contaminated site. In a further study from the same research group, Maier et al. (2013), optimized the ESIA separation of phenoxy acids by using a slower oven temperature ramp (1.5 °C/min) and a relatively high velocity (29 cm/s corresponding to a He carrier flow rate of 1.4 ml/min) and therefore inducing the separation of enantiomers at a lower temperature. Maier et al. (2013) refined the applicability of ESIA to polar herbicides such as the following phenoxy acids: 4-CPP ((RS)-2-(4-chlorophenoxy)propionic acid), mecoprop (2-(4-chloro-2-methylphenoxy)-propionic acid), and dichlorprop (2-(2,4-dichlorophenoxy)-propionic acid). The accuracy of the isotope composition measured in ESIA of the abovementioned polar herbicides was also checked through a linearity test and the authors recommended that for accurate ESIA of polar herbicides, only peak amplitudes higher than 200 mV should be used. Also the ESIA was applied to investigate the aerobic biodegradation of dichlorprop (2-(2,4-dichlorophenoxy)-propionic acid) with the pure strain of Delftia acidovorans MC1 (Muller and Hoffmann, 2006) and the experiment showed pronounced enantiomer fractionation, but no isotope fractionation. Qiu et al. (2014) provided the most extensive study published until now on ESIA of phenoxy acids by investigating (enantiomer-specific) isotope fractionation in aerobic degradation of five different phenoxy acids (two phenoxy acetic, three phenoxy propionic acids) with a comprehensive number of phenoxy acid-degrading microorganisms (Sphingobium herbicidovorans MH, Sphingomonas sp. ERG5, Sphingomonas sp. PM2, D. acidovorans MC1, Cupriavidus pinatubonensis JMP134, and Cupriavidus basilensis sp. ERG4). Small isotope enrichment factors were observed during degradation of MCPA (4-chloro-2methylphenoxyacetic acid) by Cupriavidus necator JMP143 (−1.3 ± 0.2‰) and C. necator JMP143 (−1.3 ± 0.2‰) and during degradation of 2,4-D (2,4-dichlorophenoxyacetic acid) by Sphingomonas sp. PM2 (−2.0 ± 0.3‰). Additionally, Qiu et al. (2014), suggested if the cell uptake were the rate-limiting step, enantiomer fractionation would be expected but not isotope fractionation; in contrast, if the enzyme reaction were the rate-limiting step, changes in both enantiomeric fraction and isotope ratios would be expected. To verify this hypothesis, Qiu et al. (2014) also performed experiments with a purified enzyme (αketoglutarate-dependent (R)-dichlorprop dioxygenase (RdpA)) to investigate the “intrinsic” isotope fractionation in the (enantioselective) enzyme reaction and to assess if the enzyme reaction or if uptake is a rate-limiting step in the whole cell degradation. An enzyme assay was conducted with RdpA from S. herbicidovorans MH, which specifically degrades (R)-DCPP (R-2-(2,4-dichlorophenoxy)-propionic acid), producing phenol and pyruvate. This enzyme assay degradation was accompanied by small isotope changes (about 2.5‰) in (R)-DCPP, whereas isotope values of (S)-DCPP, which were not transformed, remained constant. Small isotope enrichment factors were calculated for the degradation of (R)-DCPP in an RdpA enzyme assay (εea = −1.0 ± 0.1‰) and an even smaller fractionation in whole cell experiments of the host organism S. herbicidovorans MH (εwc = −0.3 ± 0.1‰). These values of isotope enrichment factors for (R)-DCPP suggest that (i) enzyme-associated isotope effects were already small, yet (ii) these isotope effects were further masked by active transport through the cell membrane. The study

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concluded that ESIA is a better indicator to detect aerobic biodegradation of phenoxypropionic acids in the field, comparing with CSIA or enantiomer fractionation applied separately. However, very few isotope enrichment factors were calculated for the isomer and enantiomers of the target phenoxy acids, therefore further studies are needed to exploit the potential of ESIA for assessment of the in situ transformation of phenoxy acids.

analysis of individual enantiomers was performed and no isotope enrichment factors for the enantiomers could be calculated. Therefore, further studies are needed to further develop the ESIA of phenoxyalkanoic methyl herbicides.

3.3. Galaxolide (HHCB)

Very recently, Zhang et al. (2014) studied the abiotic transformation of α-HCH by direct and indirect photolysis, alkaline hydrolysis, electrochemical reduction and by reduction with Fe0 nanoparticles. The lowest carbon isotope enrichment factors of individual were recorded for indirect photolysis under H2O2 (−1.7 ± 0.2 for εC(+) and −2.1 ± 0.3 for εC(−) respectively) and the highest for alkaline hydrolysis (− 7.2 ± 0.5 for εC(+) and −7.7 ± 0.6 for εC(−) respectively). As expected, Zhang et al. (2014) concluded that enantioselective transformation of α-HCH was observed in none of the abovementioned chemical reactions. In contrast to enantioselective biodegradation, the carbon isotope enrichment factors of individual enantiomers (εC(+), εC(−)) in all chemical transformation were statistically identical with each other and with the respective εC of bulk α-HCH within a 95% confidence interval. Zhang et al. (2014) compared also chemical and biological isotope fractionation of α-HCH. For instance, the dehydrochlorination of α-HCH by alkaline hydrolysis and aerobic biodegradation was compared. Both processes produced PCCH and TCBs as metabolites, indicating an E2 elimination reaction for both transformation pathways. In comparison to aerobic biodegradation (the bulk εC values were −1.6 ± 0.3‰ for the S. indicum strain B90A and −1.0 ± 0.2‰ for the S. japonicum strain UT26 respectively), the alkaline hydrolysis showed a much higher carbon isotope fractionation (bulk εC value was −7.6 ± 0.4‰). Similarly, dichloroelimination catalyzed by Fe0 nanoparticles via two-electron transfers gave a slightly higher carbon isotope fractionation (the bulk εC of − 4.9 ± 0.1‰) than dichloroelimination by anaerobic biodegradation (the bulk εC of − 3.7 ± 0.8‰) (Badea et al., 2011). In both cases, Zhang et al. (2014) suggested that the binding of the substrate to the enzyme during aerobic and anaerobic biodegradation may reduce the extent of carbon isotope fractionation as compared to direct bond cleavage in pure chemical reactions. By comparing the isotope fractionation between α-HCH enantiomers, further information on rate limitation associated with enzyme-substrate binding can be obtained. The difference in εC of the α-HCH enantiomers (ΔεC = 1.8‰) in the experiments with the S. indicum strain B90A, quantitatively demonstrated the masking effect on carbon isotope fractionation by binding to enzyme suggesting that binding of enantiomers to the enzyme might play a role in the overall process of isotope fractionation. Zhang et al. (2014) also tried to distinguish between abiotic and biotic transformation pathways of α-HCH using the combination of enantiomer and carbon isotope fractionation, in a two-dimensional approach developed by Bashir et al. (2013). However, they found that the range of chemical transformation partially overlaps with those of anaerobic biodegradation. Zhang et al. (2014) conclude that ESIA is a feasible concept for characterizing the transformation of α-HCH. However, due to the low enantioselectivity for both chemical transformation and anaerobic biodegradation of α-HCH, the applicability to distinguish these reactions is limited.

Wang et al. (2013) developed an ESIA for the enantiomers of galaxolide (HHCB), a synthetic polycyclic musk that is widely used in personal care products and during the last 10–15 years has increasingly raised public concern due to its bioaccumulation in biota, incompletely understood environmental fate and potentially adverse effects on human health and the environment (Bester, 2005; Franke et al., 1999; Gatermann et al., 2002; Martin et al., 2007; Peck et al., 2006; Zeng et al., 2008). Wang et al., 2013 achieved a good baseline separation for each of the four HHCB enantiomers by GC–MS analysis. The difference between the retention times for the (−)-trans-HHCB and (−)-cisHHCB enantiomers was more than 2 min, while the difference for (+)-trans-HHCB and (+)-cis-HHCB enantiomers was about 1 min. However, on the GC–C–IRMS system, nearly complete baseline separation was still observed, although the differences between the isotopic signatures of the HHCB enantiomers were only 1 and 0.7 min. The average isotope ratios of the four HHCB enantiomers from a standard solution (LGC Promochem, Weseel, Germany) were ranged from − 26.21‰ to − 26.50‰, very close to those of the HHCB racemic mixture (−26.58‰), demonstrating the accuracy of the method. Furthermore, a sediment sample was used to test the developed method, and it was shown that the HHCB enantiomers in the sediment sample exhibited significantly different isotopic signature: for the transenantiomers (H1 and H4), the δ13C values were − 33.03‰ and −32.69‰, respectively, whereas, for the cis-enantiomers (H2 and H3), the δ13C values were − 27.23‰ and − 24.57‰, respectively. For the same sediment sample, a slight enantiomeric fractionation (0.507 for (−)-trans-HHCB and 0.490 for (−)-cis-HHCB respectively) was observed. It can be concluded that (Wang et al., 2013) observed carbon isotope ratios are exhibiting significant isotopic fractionation, suggesting that both cis-HHCB enantiomers had undergone a possible biotransformation in the sediment comparing with both trans-HHCB enantiomers, while the EF values of the two pairs of HHCB enantiomers in the sediment were similar to those of the HHCB standards. However with a respect of sediment samples, according to the data provided by Wang et al., 2013 an isotope balance could not be calculated since HHCB has four enantiomers ((−)-trans-HHCB, (+)-trans-HHCB and (−)-cis-HHCB (+)-cis-HHCB) and the enantiomeric fractions of cisHHCB and trans-HHCB were calculated individually for each pair and not taking into consideration the sum of areas of all four peaks, while the isotope composition of the bulk HHCB was determined jointly for all four enantiomers. Thus, further studies are needed to exploit the potential of ESIA for the enantiomers of galaxolide (HHCB) in environment. 3.4. Phenoxyalkanoic methyl herbicides Jammer et al. (2014) studied the enantioselective enzymatic hydrolysis of mecoprop-methyl, dichlorprop-methyl, and dimethylmethylsuccinate by lipases from Pseudomonas fluorescens, Pseudomonas cepacia, and Candida rugosa, and both isotope and enantiomeric enrichment factors were calculated. The highest isotope environment factor was obtained for the bulk (R,S)-dimethyl 2-methylsuccinate (DMMS) during enzymatic hydrolysis by lipases extracted from C. rugosa (−3.4 ± 0.5) and the highest enantiomeric enrichment factor for enzymatic hydrolysis of methyl 2-(4-chloro-2-methylphenoxy) propanoate (MCPPM) by lipase from C. rogusa (−194.1 ± 4.4). However, no isotope

4. ESIA as a differential tool between abiotic and biotic transformation of organic contaminants

5. Conclusions The abovementioned studies demonstrated that ESIA is an innovative tool for the characterization of the environmental fate of chiral organic contaminants and was applied until now only to a few chiral organic contaminants. However, as shown earlier, the ESIA requires a good separation between the enantiomers on chiral GC columns. The development of new commercially available chiral GC column might extend the applicability of ESIA to more chiral emerging contaminants. Furthermore, since many chiral organic contaminants of high molecular

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mass cannot be determined by chiral GC analysis, further development of ESIA using liquid chromatography–combustion–isotope ratio mass spectrometer (LC–C–IRMS) systems is expecting to appear in the following years. The abovementioned study has shown that ESIA can be used to distinguish between biotic and abiotic degradation pathways of chiral organic contaminants. However, the distinction of these transformation processes seems to be not always possible when the degradation of organic contaminants is accompanied not only by a significant isotope fractionation but also by a low or negligible enantiomeric fractionation. In the future, it is expected that a combination between ESIA and a multi-element isotope analysis (i.e. hydrogen vs. carbon vs. chorine) is to be developed for a better characterization of transformation pathways of chiral organic contaminants.

References Amaral, H.I.F., Berg, M., Brennwald, M.S., Hofer, M., Kipfer, R., 2010. C-13/C-12 analysis of ultra-trace amounts of volatile organic contaminants in groundwater by vacuum extraction. Environ. Sci. Technol. 44, 1023–1029. Badea, S.L., Vogt, C., Weber, S., Danet, A.F., Richnow, H.H., 2009. Stable isotope fractionation of gamma-hexachlorocyclohexane (lindane) during reductive dechlorination by two strains of sulfate-reducing bacteria. Environ. Sci. Technol. 43, 3155–3161. Badea, S.L., Vogt, C., Gehre, M., Fischer, A., Danet, A.F., Richnow, H.H., 2011. Development of an enantiomer-specific stable carbon isotope analysis (ESIA) method for assessing the fate of alpha-hexachlorocyclohexane in the environment. Rapid Commun. Mass Spectrom. 25, 1363–1372. Bashir, S., Fischer, A., Nijenhuis, I., Richnow, H.H., 2013. Enantioselective carbon stable isotope fractionation of hexachlorocyclohexane during aerobic biodegradation by Sphingobium spp. Environ. Sci. Technol. 47 (20), 11432–11439. Berg, M., Bolotin, J., Hofstetter, T.B., 2007. Compound-specific nitrogen and carbon isotope analysis of nitroaromatic compounds in aqueous samples using solid-phase microextraction coupled to GC/IRMS. Anal. Chem. 79, 2386–2393. Bester, K., 2005. Polycyclic musks in the Ruhr catchment area — transport, discharges of waste water, and transformations of HHCB, AHTN and HHCB-lactone. J. Environ. Monit. 7, 43–51. Buser, H.R., Muller, M.D., 1995. Isomer and enantioselective degradation of hexachlorocyclohexane isomers in sewage-sludge under anaerobic conditions. Environ. Sci. Technol. 29, 664–672. Cichocka, D., Siegert, M., Imfeld, G., Andert, J., Beck, K., Diekert, G., et al., 2007. Factors controlling the carbon isotope fractionation of tetra- and trichloroethene during reductive dechlorination by Sulfurospirillum ssp and Desulfitobacterium sp. strain PCE-S. FEMS Microbiol. Ecol. 62, 98–107. Coplen, T., Brand, W., Gehre, M., Groning, M., Meijer, H., Toman, B., et al., 2006. After two decades a second anchor for the VPDB delta C-13 scale. Rapid Commun. Mass Spectrom. 20, 3165–3166. de Geus, H.J., Wester, P.G., de Boer, J., Brinkman, U.A.T., 2000. Enantiomer fractions instead of enantiomer ratios. Chemosphere 41, 725–727. Drenzek, N., Eglinton, T., May, J., Wu, Q., Sowers, K., Reddy, C., 2001. The absence and application of stable carbon isotopic fractionation during the reductive dechlorination of polychlorinated biphenyls. Environ. Sci. Technol. 35, 3310–3313. Eljarrat, E., Guerra, P., Barcelo, D., 2008. Enantiomeric determination of chiral persistent organic pollutants and their metabolites. Trac-Trends in Anal Chem 27, 847–861. Elsner, M., 2010. Stable isotope fractionation to investigate natural transformation mechanisms of organic contaminants: principles, prospects and limitations. J. Environ. Monit. 12, 2005–2031. Elsner, M., Zwank, L., Hunkeler, D., Schwarzenbach, R.P., 2005. A new concept linking observable stable isotope fractionation to transformation pathways of organic pollutants. Environ. Sci. Technol. 39, 6896–6916. Elsner, M., Jochmann, M.A., Hofstetter, T.B., Hunkeler, D., Bernstein, A., Schmidt, T.C., et al., 2012. Current challenges in compound-specific stable isotope analysis of environmental organic contaminants. Anal. Bioanal. Chem. 403, 2471–2491. Fischer, A., Herklotz, I., Herrmann, S., Thullner, M., Weelink, S.A.B., Stams, A.J.M., et al., 2008. Combined carbon and hydrogen isotope fractionation investigations for elucidating benzene biodegradation pathways. Environ. Sci. Technol. 42, 4356–4363. Franke, S., Meyer, C., Heinzel, N., Gatermann, R., Huhnerfuss, H., Rimkus, G., et al., 1999. Enantiomeric composition of the polycyclic musks HHCB and AHTN in different aquatic species. Chirality 11, 795–801. Gasser, G., Pankratov, I., Elhanany, S., Werner, P., Gun, J., Gelman, F., et al., 2012. Field and laboratory studies of the fate and enantiomeric enrichment of venlafaxine and Odesmethylvenlafaxine under aerobic and anaerobic conditions. Chemosphere 88, 98–105. Gatermann, R., Biselli, S., Huhnerfuss, H., Rimkus, G.G., Franke, S., Hecker, M., et al., 2002. Synthetic musks in the environment. Part 2: enantioselective transformation of the polycyclic musk fragrances HHCB, AHTN, AHDI, and ATII in freshwater fish. Arch. Environ. Contam. Toxicol. 42, 447–453. Hardt, I.H., Wolf, C., Gehrcke, B., Hochmuth, D.H., Pfaffenberger, B., Huhnerfuss, H., et al., 1994. Gas chromatographic enantiomer separation of agrochemicals and polychlorinated biphenyls (PCBs) using modified cyclodextrins. J. High Resolut. Chromatogr. 17, 859–864.

465

Harner, T., Wiberg, K., Norstrom, R., 2000. Enantiomer fractions are preferred to enantiomer ratios for describing chiral signatures in environmental analysis. Environ. Sci. Technol. 34, 218–220. Harrison, I., Leader, R.U., Higgo, J.J.W., Williams, G.M., 1998. A study of the degradation of phenoxy acid herbicides at different sites in a limestone aquifer. Chemosphere 36, 1211–1232. Horii, Y., Kannan, K., Petrick, G., Gamo, T., Falandysz, J., Yamashita, N., 2005. Congenerspecific carbon isotopic analysis of technical PCB and PCN mixtures using twodimensional gas chromatography–isotope ratio mass spectrometry. Environ. Sci. Technol. 39, 4206–4212. Horii, Y., van Bavel, B., Kannan, K., Petrick, G., Nachtigall, K., Yamashita, N., 2008. Novel evidence for natural formation of dioxins in ball clay. Chemosphere 70, 1280–1289. Jammer, S., Voloshenko, A., Gelman, F., Lev, O., 2014. Chiral and isotope analyses for assessing the degradation of organic contaminants in the environment: Rayleigh dependence. Environ. Sci. Technol. 48, 3310–3318. la Farre, M., Perez, S., Kantiani, L., Barcelo, D., 2008. Fate and toxicity of emerging pollutants, their metabolites and transformation products in the aquatic environment. TrAC Trends Anal. Chem. 27, 991–1007. Liu, F., Cichocka, D., Nijenhuis, I., Richnow, H.H., Fennell, D.E., 2010. Carbon isotope fractionation during dechlorination of 1,2,3,4-tetrachlorodibenzo-p-dioxin by a Dehalococcoides-containing culture. Chemosphere 80, 1113–1119. Lollar, B., Slater, G., Sleep, B., Witt, M., Klecka, G., Harkness, M., et al., 2001. Stable carbon isotope evidence for intrinsic bioremediation of tetrachloroethene and trichloroethene at area 6, Dover Air Force Base. Environ. Sci. Technol. 35, 261–269. Maier, M.P., Qiu, S.R., Elsner, M., 2013. Enantioselective stable isotope analysis (ESIA) of polar herbicides. Anal. Bioanal. Chem. 405, 2825–2831. Marco-Urrea, E., Nijenhuis, I., Adrian, L., 2011. Transformation and carbon isotope fractionation of tetra- and trichloroethene to trans-dichloroethene by Dehalococcoides sp. strain CBDB1. Environ. Sci. Technol. 45, 1555–1562. Martin, C., Moeder, M., Daniel, X., Krauss, G., Schlosser, D., 2007. Biotransformation of the polycyclic musks HHCB and AHTN and metabolite formation by fungi occurring in freshwater environments. Environ. Sci. Technol. 41, 5395–5402. Meckenstock, R.U., Morasch, B., Griebler, C., Richnow, H.H., 2004. Stable isotope fractionation analysis as a tool to monitor biodegradation in contaminated aquifers. J. Contam. Hydrol. 75, 215–255. Meier-Augenstein, W., Watt, P.W., Langhans, C.D., 1996. Influence of gas chromatographic parameters on measurement of 13C/12C isotope ratios by gas–liquid chromatography-combustion isotope ratio mass spectrometry. J. Chromatogr. A 752, 233–241. Meyer, A.H., Penning, H., Lowag, H., Elsner, M., 2008. Precise and accurate compound specific carbon and nitrogen isotope analysis of atrazine: critical role of combustion oven conditions. Environ. Sci. Technol. 42, 7757–7763. Milosevic, N., Qiu, S., Elsner, M., Einsiedl, F., Maier, M.P., Bensch, H.K.V., et al., 2013. Combined isotope and enantiomer analysis to assess the fate of phenoxy acids in a heterogeneous geologic setting at an old landfill. Water Res. 47, 637–649. Muller, R.H., Hoffmann, D., 2006. Uptake kinetics of 2,4-dichlorophenoxyacetate by Delftia acidovorans MC1 and derivative strains: complex characteristics in response to pH and growth substrate. Biosci. Biotechnol. Biochem. 70, 1642–1654. Muller, T.A., Kohler, H.P.E., 2004. Chirality of pollutants — effects on metabolism and fate. Appl. Microbiol. Biotechnol. 64, 300–316. Muller, M.D., Schlabach, M., Oehme, M., 1992. Fast and precise determination of alphahexachlorocyclohexane enantiomers in environmental-samples using chiral highresolution gas-chromatography. Environ. Sci. Technol. 26, 566–569. Nijenhuis, I., Andert, J., Beck, K., Kastner, M., Diekert, G., Richnow, H., 2005. Stable isotope fractionation of tetrachloroethene during reductive dechlorination by Sulfurospirillum multivorans and Desulfitobacterium sp. strain PCE-S and abiotic reactions with cyanocobalamin. Appl. Environ. Microbiol. 71, 3413–3419. Oulhote, Y., Le Bot, B., Deguen, S., Glorennec, P., 2011. Using and interpreting isotope data for source identification. TrAC Trends Anal. Chem. 30, 302–312. Padma, T.V., Dickhut, R.M., Ducklow, H., 2003a. Variations in alpha-hexachlorocyclohexane enantiomer ratios in relation to microbial activity in a temperate estuary. Environ. Toxicol. Chem. 22, 1421–1427. Padma, T.V., Dickhut, R.M., Ducklow, H., 2003b. Variations in α-hexachlorocyclohexane enantiomer ratios in relation to microbial activity in a temperate estuary. Environ. Toxicol. Chem. 22, 1421–1427. Patterson, D., Schnell, M., 2014. New studies on molecular chirality in the gas phase: enantiomer differentiation and determination of enantiomeric excess. Phys. Chem. Chem. Phys. 16, 11114–11123. Peck, A.M., Linebaugh, E.K., Hornbuckle, K.C., 2006. Synthetic musk fragrances in Lake Erie and Lake Ontario sediment cores. Environ. Sci. Technol. 40, 5629–5635. Penning, H., Elsner, M., 2007. Intramolecular carbon and nitrogen isotope analysis by quantitative dry fragmentation of the phenylurea herbicide isoproturon in a combined injector/capillary reactor prior to GC separation. Anal. Chem. 79, 8399–8405. Qiu, S.R., Gozdereliler, E., Weyrauch, P., Lopez, E.C.M., Kohler, H.P.E., Sorensen, S.R., et al., 2014. Small C-13/C-12 fractionation contrasts with large enantiomer fractionation in aerobic biodegradation of phenoxy acids. Environ. Sci. Technol. 48, 5501–5511. Reddy, C., Drenzek, N., Eglinton, T., Heraty, L., Sturchio, N., Shiner, V., 2002. Stable chlorine intramolecular kinetic isotope effects from the abiotic dehydrochlorination of DDT. Environ Sci Pol Res 9, 183–186. Rosell, M., Barcelo, D., Rohwerder, T., Breuer, U., Gehre, M., Richnow, H.H., 2007. Variations in C-13/C-12 and D/H enrichment factors of aerobic bacterial fuel oxygenate degradation. Environ. Sci. Technol. 41, 2036–2043. Schurig, V., 1994. Enantiomer separation by gas-chromatography on chiral stationary phases. J. Chromatogr. A 666, 111–129.

466

S.-L. Badea, A.-F. Danet / Science of the Total Environment 514 (2015) 459–466

Shi, X.Y., Liang, P., Gao, X.W., 2006. Determination enantiomer excess (e.e. %) of chiral sharpless epoxides with beta-cyclodextrin derivatives as chiral stationary phases of capillary gas chromatography. Chin. Chem. Lett. 17, 505–508. Suar, M., Hauser, A., Poiger, T., Buser, H.R., Muller, M.D., Dogra, C., et al., 2005. Enantioselective transformation of alpha-hexachlorocyclohexane by the dehydrochlorinases LinA1 and LinA2 from the soil bacterium Sphingomonas paucimobilis B90A. Appl. Environ. Microbiol. 71, 8514–8518. Vetter, W., Gleixner, G., Armbruster, W., Ruppe, S., Stern, G.A., Braekevelt, E., 2005. Congener-specific concentrations and carbon stable isotope ratios (δ13C) of two technical toxaphene products (Toxaphene® and Melipax®). Chemosphere 58, 235–241. Vetter, W., Gaul, S., Armbruster, W., 2008. Stable carbon isotope ratios of POPS — a tracer that can lead to the origins of pollution. Environ. Int. 34, 357–362. Wang, J., Gao, S., Zeng, X., Yu, Z., Pa, Peng, Sheng, G., et al., 2013. Compound-specific stable carbon isotope analysis of galaxolide enantiomers in sediment using gas chromatography/isotope ratio monitoring mass spectrometry. Rapid Commun. Mass Spectrom. 27, 1690–1696. Wiberg, K., Brorstrom-Lunden, E., Wangberg, I., Bidleman, T.F., Haglund, P., 2001a. Concentrations and fluxes of hexachlorocyclohexanes and chiral composition of alphaHCH in environmental samples from the southern Baltic Sea. Environ. Sci. Technol. 35, 4739–4746.

Wiberg, K., Harner, T., Wideman, J.L., Bidleman, T.F., 2001b. Chiral analysis of organochlorine pesticides in Alabama soils. Chemosphere 45, 843–848. Zeng, X.Y., Mai, B.X., Sheng, G.Y., Luo, X.J., Shao, W.L., An, T.C., et al., 2008. Distribution of polycyclic musks in surface sediments from the Pearl River Delta and Macao coastal region, South China. Environ. Toxicol. Chem. 27, 18–23. Zhang, N., Bashir, S., Qin, J., Schindelka, J., Fischer, A., Nijenhuis, I., et al., 2014. Compound specific stable isotope analysis (CSIA) to characterize transformation mechanisms of α-hexachlorocyclohexane. J. Hazard. Mater. 280, 750–757. Zipper, C., Bunk, M., Zehnder, A.J.B., Kohler, H.P.E., 1998a. Enantioselective uptake and degradation of the chiral herbicide dichlorprop (RS)-2-(2,4-dichlorophenoxy)propanoic acid by Sphingomonas herbicidovorans MH. J. Bacteriol. 180, 3368–3374. Zipper, C., Suter, M.J.F., Haderlein, S.B., Gruhl, M., Kohler, H.P.E., 1998b. Changes in the enantiomeric ratio of (R)- to (S)-mecoprop indicate in situ biodegradation of this chiral herbicide in a polluted aquifer. Environ. Sci. Technol. 32, 2070–2076. Zipper, C., Fleischmann, T., Kohler, H.P.E., 1999. Aerobic biodegradation of chiral phenoxyalkanoic acid derivatives during incubations with activated sludge. FEMS Microbiol. Ecol. 29, 197–204.

Enantioselective stable isotope analysis (ESIA) - a new concept to evaluate the environmental fate of chiral organic contaminants.

Since 2011, the enantiospecific stable carbon isotope analysis (ESIA) has emerged as an innovative technique to assess the environmental fate of chira...
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