Compound Specific and Enantioselective Stable Isotope Analysis as

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Compound Specific and Enantioselective Stable Isotope Analysis as tools to monitor transformation of hexachlorocyclohexane (HCH) in a complex aquifer system Yaqing Liu, Safdar Bashir, Reiner Stollberg, Ralf Trabitzsch, Holger Weiss, Heidrun Paschke, Ivonne Nijenhuis, and Hans Hermann Richnow Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b05632 • Publication Date (Web): 04 Jul 2017 Downloaded from http://pubs.acs.org on July 4, 2017

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Environmental Science & Technology

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Compound Specific and Enantioselective Stable Isotope Analysis as tools to monitor

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transformation of hexachlorocyclohexane (HCH) in a complex aquifer system

3

YAQING LIU1†, SAFDAR BASHIR1†#, REINER STOLLBERG2, RALF TRABITZSCH2,

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HOLGER WEIß², HEIDRUN PASCHKE3, IVONNE NIJENHUIS1*, HANS-HERMANN

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RICHNOW1.

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1

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Permoserstraße 15, 04318 Leipzig, Germany

8

2

9


Department of Isotope Biogeochemistry, Helmholtz Centre for Environmental Research-UFZ,

Department Groundwater Remediation, Helmholtz Centre for Environmental Research-UFZ,

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3

11


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#

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Faisalabad 38040, Pakistan

14 15

†

Department of Analytical Chemistry, Helmholtz Centre for Environmental Research-UFZ,

Current address: Institute of Soil & Environmental Sciences, University of Agriculture,

These authors contributed equally to this work

*Corresponding author: Ivonne Nijenhuis

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1 ACS Paragon Plus Environment


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ABSTRACT

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Technical hexachlorocyclohexane (HCH) mixtures and Lindane (-HCH) have been produced in

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Bitterfeld-Wolfen, Germany, for about 30 years until 1982. In the vicinity of the former dump

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sites and production facilities, large plumes of HCHs persist within two aquifer systems. We

21

studied the natural attenuation of HCH in these groundwater systems through a combination of

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enantiomeric and carbon isotope fractionation in order to characterize the degradation of α-HCH

23

in the areas downstream of a former disposal and production site in Bitterfeld-Wolfen. The

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concentration and isotope composition of α-HCH from the Quaternary and Tertiary aquifers

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were analyzed. The carbon isotope compositions were compared to the source signal of waste

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deposits for the dumpsite and highly contaminated areas. The average value of δ13C at dumpsite

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was -29.7 ± 0.3 ‰ and -29.0 ± 0.1 ‰ for (-) and (+)α-HCH, respectively, while those for the β-,

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γ-, δ-HCH isomers were -29.0 ± 0.3 ‰, -29.5 ± 0.4 ‰, -28.2 ± 0.2 ‰, respectively. In the

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plume, the enantiomer fraction shifted up to 0.35, from 0.50 at source area to 0.15 (well T1), and

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was found accompanied by a carbon isotope enrichment of 5 ‰ and 2.9 ‰ for (-) and (+)α-

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HCH, respectively. The established model for interpreting isotope and enantiomer fractionation

32

patterns showed potential for analyzing the degradation process at a field site with a complex

33

history with respect to contamination and fluctuating geochemical conditions.

34

KEYWORDS

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Compound-specific Stable Isotope Analysis, Enantiomer-specific Stable Isotope Analysis,

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Enantiomer fractionation, hexachlorocyclohexane, groundwater, biodegradation


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INTRODUCTION

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Hexachlorocyclohexane (HCH) isomers are pollutants of particular concern because of their

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widespread distribution in the environment, toxicity and persistence

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(Lindane) has a specific pesticide activity, the purification of Lindane resulted in the production

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of other waste residues, so-called ‘HCH muck’, which were mostly dumped near the production

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site. The widespread application of Lindane and a large amount of the other HCH isomers as by-

43

products have caused contamination in soil, groundwater and atmosphere 4-9.

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HCH is biodegradable under both oxic and anoxic conditions

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which were able to degrade HCHs have been reported

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hydrolysis has been considered as the dominant abiotic transformation for HCH in field sites

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although associated with very long half life time at acidic and neutral conditions

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Investigations on bioremediation of HCH contaminated soil and groundwater have already been

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conducted in laboratory and field scale 14-17.

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Within contaminated aquifers, monitoring of concentration levels alone does not allow

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identifying degradation processes as the decrease in concentration may be due to physical

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processes such as volatilization, sorption, dilution and dispersion. Compound-specific stable

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isotope analysis (CSIA) may allow distinguishing degradation processes from non-destructive

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processes

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energy for bond cleavage and thus tend to be degraded faster than molecules containing the

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heavy carbon isotope (13C), resulting in a

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pollutant. CSIA has been applied to monitor in situ biodegradation of a wide variety of organic

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contaminants in aquifers

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differentiation and identification of in situ biodegradation in the groundwater of an operating

18, 19

11

10-12

1-3

. Since only γ-HCH

and several microorganisms

. For chemical degradation, alkaline

13

.

. Molecules with light carbon isotopes (12C) in the reactive position require less

20

13

C-enrichment in the remaining fraction of the

. Thus far, CSIA of HCH isomers showed potential for HCH source



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21

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packaging and reformulating pesticide facility located in northeastern Florida, USA

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CSIA was applied for the assessment of HCH natural attenuation processes within contaminated

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aquifers in the area of a former pesticide formulating plant in Germany 22. Furthermore, triple-

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element (H, C, Cl) stable isotope analysis facilitated source identification of HCH products, and

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may also be used to assess transformation processes 23.

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CSIA of HCH can be used to characterize the degradation processes including both biological

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and chemical transformation

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provides an indicator for biodegradation of chiral compounds

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isomer of the eight HCH isomers and the enantiomer specific biodegradation of α-HCH results in

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enrichment of one enantiomer in the non-degraded residual phase which leads to changes in the

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enantiomeric fraction

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groundwater has been described as an indicator for biodegradation in the field

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enantiomer specific transformation of α-HCH was observed for biotic (aerobic and anaerobic)

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transformation, however no enantiomer specific transformation was observed in case of chemical

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transformation 11, 29, 33-35.

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The combination of CSIA and EF for the investigation of HCH degradation was done so far in

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laboratory studies

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isotope fractionation for chiral pesticides 36, but to our best knowledge not for the evaluation of

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in situ biodegradation at a field scale. Therefore, in this study, compound-specific and

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enantiomer-specific stable isotope analysis (ESIA) were combined with the analysis of

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enantiomer fractionation and applied at a field site to analyze transformation pathways of α-

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HCH. Hydrogeochemical parameters and pollutant concentration levels, carbon isotope ratios of

34, 35

24, 25

. Also,

. Furthermore, change in the enantiomeric fraction (EF) 26-28

. α-HCH is the only chiral

29

. The enantiomer specific degradation of α-HCH in air, soil and 28, 30-32

and

, and a modeling study for joint interpretation of enantiomer and stable


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HCHs and enantiomer fraction of -HCH were determined during two groundwater monitoring

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campaigns conducted in the Bitterfeld-Wolfen area, Germany, in 2012 and 2014. A model was

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established for interpreting fractionation patterns at a field site with a complex history with

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respect to contamination and fluctuating geochemical conditions.

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MATERIALS AND METHODS

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Field site

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The Bitterfeld-Wolfen region being located in Eastern Germany, former German Democratic

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Republic territory, had been heavily impacted by open-pit lignite mining and related carbon-

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based chemical industry for more than a century until the German reunification in 1989/90

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Beside the industrial manufacture of about 4500 chlorine-based chemical substances or

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associated consumer goods, HCH and DDT were extensively synthesized in Bitterfeld-Wolfen

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between 1951 – 1982 36. Numerous former open-pit mines were subsequently used for dumping

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chemical residues from industrial production without any appropriate safety or environmental

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protection measures. The most relevant landfill regarding its toxic inventory is named “Antonie”

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(Figure 1 – overview map), which is containing in total about 5,000,000 m³ of various chemical

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compounds while about 70,000 tons of the landfilling are estimated to be HCH isomers and had

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been dumped between 1962 and 1982

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from HCH production sites, storage and loading areas (‘Area C’, Figure 1) caused a multi-source

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environmental pollution at the regional scale which affected all surrounding environmental

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compartments such as air, soil, nearby surface waters and in particular the regional groundwater

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system

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contaminated by unhindered HCH release from surrounding waste disposals (Figure 1, Figure 2)

39, 40

37, 38

37

.

. Leaking waste dumps as well as contaminant inputs

. Groundwater volumes of the regional lower and upper aquifer are heavily



40

104

at an area of about 30-35 square kilometers

. Additional detailed information about the

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contaminant history and its complexity as well as the regional hydrogeological setting of the

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Bitterfeld-Wolfen region can be found in the Supplementary Material (SI-S1).

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Figure 1 （p23）

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Figure 2 （p24）

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Sampling

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Sampling campaigns were performed in 2012 and 2014, during which 42 groundwater wells

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were sampled and three ‘HCH muck’ samples were collected at two dumpsites. For CSIA,

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groundwater was filled in three one L glass bottles (Schott, Germany) sealed with Teflon-coated

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caps (Schott, Germany) without headspace, thus avoiding evaporation. The water samples were

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adjusted to pH 2 using hydrochloric acid (HCl; 25%, Carl Roth GmbH & Co. KG, Germany) to

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inhibit microbial activity. Standard sampling methods can be found in supporting information SI-

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S2.

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HCH muck samples were taken from area C (one sample) and from the dump site Antonie (two

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samples). These were considered to represent the isotope and enantiomeric composition of the

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original waste product showing a racemic distribution of α-HCH enantiomers. The muck

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contained grey to white crystals made up of > 90% weight percent of HCH isomers. The muck

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was almost completely soluble in n-hexane. The isomer distribution of α-, β-, γ- and δ-HCH was

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approximate to 85%, 10%, 1% and 0.4%, respectively. We assumed that these HCHs were not

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significantly affected by biodegradation and represented the original mixture of waste material

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from HCH production.


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Analytical procedures

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Concentration analysis: Concentrations of dissolved oxygen, temperature, pH, redox potential,

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and electrical conductivity were determined during sampling using appropriate electrodes using a

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Multimeter from WTW GmbH, Germany, equipped with dissolved oxygen sensor (CellOx®

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325), pH electrode system (SenTix® 41), redox electrode (SenTix® ORP) and a conductivity

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cell (KLE 325), see supporting information SI-S2. For the sampling campaign of 2012, samples

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for concentration analyses of pollutants and hydrochemical parameters were immediately

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processed as described in detail in the SI-S2. In 2014, concentration analyses of pollutants and

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hydrochemical parameters were determined by professional analytical companies. For the

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methods details please see supporting information SI-S2. Data on concentration and geochemical

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parameters were provided by the Landesanstalt für Altlastenfreistellung (LAF).

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Compound-specific Stable Isotope Analysis: Samples for isotope analysis were stored in the

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dark at 4 °C until extraction. One L water sample was mixed with 30 mL dichloromethane

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(DCM; ≥ 99.8%, Carl Roth GmbH & Co. KG, Germany) in a 1L separating funnel, and was

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repeated twice for each sample. The DCM extracts from the same sampling well were combined

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and dried with anhydrous sodium sulfate (Na2SO4; ≥99%, Bernd Kraft GmbH, Germany) 22. An

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evaporator (TurboVap® II, Biotage AB, Sweden) was used to concentrate the samples to one mL

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for concentration and carbon isotope analysis of HCHs.

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The isotope composition of HCHs was analyzed by gas chromatography isotope ratio mass

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spectrometry (GC-IRMS), as described previously 33. Quality control was done by using isotope

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laboratory standards consisting of α-HCH (99%, Sigma-Aldrich Chemie GmbH, Germany) with

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carbon isotope ratios determined by elemental analyzer isotope ratio mass spectrometry (EA7 ACS Paragon Plus Environment


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IRMS). The carbon isotope ratios of pure α-HCH measured by GC-IRMS were reported in the δ

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notation (δ13C) relative to the international standard Vienna Pee Dee Belemnite (VPDB)

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according to eq.1 39.

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𝛿 13 𝐶𝑠𝑎𝑚𝑝𝑙𝑒 = 𝑅

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Rsample and Rstandard are the 13C/12C ratios of the samples and VPDB, respectively. The δ13C-values

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were reported in per mil (‰). All the samples were measured in triplicate.

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Enantiomer fractionation: The enantiomeric ratio was analyzed using a GC- IRMS for samples

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with a concentration higher than 1.0 μg L-1 and GC-MS for lower ones (for detailed information

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please see SI-S3) equipped with γ-DEX 120 chiral column (Sigma-Aldrich). The EF(+) is

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defined as A+/(A++A-) and EF(-) is defined as A-/ (A++A-), where A+ and A- correspond to the

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peak areas or concentrations of (+) and (-) enantiomers 40. An EF (+) > 0.5 shows the preferential

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degradation of (-) enantiomer, and an EF (+) < 0.5 indicates the preferential degradation of (+)

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enantiomer.

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Estimation of biodegradation: In this study, we set a model for the evaluation of the

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degradation through isotope fractionation analysis and enantiomer fractionation analysis. For the

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quantification of α-HCH degradation by isotope analysis, the Rayleigh equation was simplified

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and applied for calculation biodegradation (BISO%), as shown in eq. 2 20.

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BISO % = [1 − (𝛿𝑡 +1)𝜀𝑐 ]*100

𝑅𝑠𝑎𝑚𝑝𝑙𝑒 𝑠𝑡𝑎𝑛𝑑𝑎𝑟𝑑

𝛿 +1 0

−1

(1)

1

(2)


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εc is the carbon isotope enrichment factor. For the calculation of α-HCH degradation, the

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enrichment factors for aerobic biodegradation were taken from Bashir et al 35which were εc1=-3.3

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±0.8 ‰ and εc2=-2.4 ±0.8 ‰.

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As significant enantiomer fractionation was observed in the samples, calculation of

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biodegradation based on enantiomer fractionation via the Rayleigh equation was also attempted.

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However, estimation of biodegradation using EF is severely limited as discussed below.

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The Rayleigh equation was also used to calculate enantiomer fractionation factors (εe), obtained

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as the slope of the linear regression line of the natural logarithm of the enantiomeric enrichment

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(ERt/ER0), against the natural logarithm of the extent of degradation eq. 3 41.

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ln 𝐸𝑅𝑡 = 𝜀𝐸𝑅 × ln 𝑓

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ER is the ratio between the more abundant to the less abundant enantiomer. In this study, ER=A-

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/A+. 𝑓 is the residual fraction (Ct/C0).

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For the quantification of α-HCH biodegradation using enantiomeric fractions (BEF (+/-)%), Eq. 4

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can be applied.

𝐸𝑅

0

𝐸𝐹(±)𝑡

(3)

1

𝐸𝑅 𝜀 (𝐸𝑅 𝑡 ) 𝑒 ]* 0

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BEF(±) % = 𝐸𝐹(±) ∗ [1 −

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Previously reported εe were applied from aerobic degradation with Sphingobium indicum strain

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B90A and Sphingobium japonicum strain UT26 for α-HCH 35.

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RESULTS AND DISCUSSION

0

100 (4)



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Hydrogeochemical parameters and conditions: The hydrogeochemical data are summarized in

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SI table S1. The geochemical conditions of the Quaternary and Tertiary aquifer are mainly

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affected by the Bitterfeld-Wolfen’s industrial history and related deposition of chemical residues.

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An acidic contamination plume with pH as low as 3.34 (well T3) stretched into the Tertiary

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aquifer from its source zone (Antonie dumpsite) towards the southeast. In relation to the

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landfill’s location, the southeastern orientation characterized its predominant groundwater

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downstream direction until the predominant hydraulic setting was changed by a regional flood

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event in August 2002. Before this century flood event, the regional groundwater level was

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significantly lower due to mining-related dewatering activities. In consequence of changing

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groundwater levels and related flow directions over the past two decades, geochemical

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conditions have changed, varied spatially and spreading of HCH from dumpsite has formed very

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complex contaminant distribution patterns. Today, the Quaternary aquifer has predominately

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neutral to slightly acidic conditions (pH was in the range of 6.3-7.4, except for monitoring well

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Q2, pH=3.79). Both aquifers comprised anoxic zones characterized by concentrations of O2

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below the detection limit (0.1 mg L-1) and elevated Fe2+ concentrations. Sulfate was present in

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both aquifers due to oxidation processes of sulfide mineral, high sulfide content indicated that

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parts of the aquifer system were sulfidogenic. Some parts of the aquifers are oxic due to long-

200

term mining-related dewatering activities over the past century, as indicated by the concentration

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of O2 (up to 0.7 mg L-1) in several wells, e.g. monitoring well Q2, Q5, T2 and T3.

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HCH concentrations: The concentrations of the four main HCH isomers are shown in Table1.

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α-HCH was detected in most of the wells except for well Q3 and Q6. In the Quaternary aquifer,

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the concentration of α-HCH at most sampling wells was < 10 µg L-1. Only in well Q2, which was

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near one of the former production sites, the concentrations of α-, β-, γ- and δ-HCH were 242 µg 10 ACS Paragon Plus Environment

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L-1, 44 µg L-1, 236 µg L-1, and 264 µg L-1, respectively. In the Tertiary aquifer, compared to

207

other wells, the higher concentration of α-HCH at well T3 and T4 were 19.8 µg L-1 and 15.6 µg

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L-1. For the other samples from this aquifer, the concentrations of α-HCH were less than 5 µg L-

209

1

210

4.5 µg L-1 (Well T4), 20.8 µg L-1 (Well T3) and 31.6 µg L-1 (Well T3), respectively.

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Table 1(p22)

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Carbon isotope fractionation of HCHs: The average carbon isotope composition of the source

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material was calculated from the isotope composition of the three analyzed muck samples

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obtained from the Antonie dumpsite and area C (Figure 1). α-HCH had a bulk carbon isotope

215

composition of –28.8 ± 1.2 ‰, with –29.5 ± 0.4 ‰ and –28.4 ± 1.0 ‰ for (-)α-HCH and (+)α-

216

HCH, respectively. The average isotope composition of β-, γ- and δ-HCH of the source material

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were –28.4 ± 1.3 ‰, –28.9 ± 1.2 ‰ and –28.2 ± 1.4 ‰, respectively (SI-table S2). The carbon

218

isotope composition of α-HCH in the three muck samples was between –27.5 ± 0.2 ‰ to –29.8 ±

219

0.3 ‰, the relatively narrow range suggesting that the carbon isotope signature was relatively

220

stable in the production process 23.

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The isotope values of α-HCH in groundwater samples were enriched in most cases compared to

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the muck samples (Figure 3). β -HCH shows an enrichment of carbon isotope in the aquifer up to

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8.6 ‰ (Well Q1) compared to the source suggesting degradation of β-HCH. γ-HCH was

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enriched up to 6.5 ‰ (Well T2) and δ-HCH up to 5.3 ‰ (Well T2) suggesting a transformation

225

in the aquifer system.

226

Figure 3 (p25)

227

In the Quaternary aquifer, the 13C signature of the (+)α-HCH and (-)α-HCH ranged from -27.6 ±

228

0.9 ‰ (Well 2012_Q7) to -23.3 ± 0.4 ‰ (Well Q8) and from -26.9 ± 0.6 ‰ (Well Q2) to -23.3 ±

, and most of them were less than 1 µg L-1. The highest concentration of β-, γ- and δ-HCH was



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0.3 ‰ (Well Q1), respectively. The resulting Δδ13C were up to 5.1 ‰ and 6.2 ‰ for (+)α-HCH

230

and (-)α-HCH, respectively, compared to the average value of the sources (SI-table S2). In the

231

Tertiary aquifer, the carbon isotope composition of (+)α-HCH and (-)α-HCH varied from -27.5 ±

232

0.2 ‰ (Well T4) to -23.4 ± 0.4 ‰ (Well T3) and from -28.6 ± 0.4 ‰ (Well T4) to -22.4 ± 0.3 ‰

233

(Well T8), respectively. Changes in δ13C values up to 5.0 ‰ and 7.1 ‰ were observed for (+)α-

234

HCH and (-)α-HCH, respectively. Additionally, for (+)α-HCH, the more significant enrichment

235

was observed for sampling wells Q8 (-23.3 ± 0.4 ‰), T3 (-23.4 ± 0.4 ‰), T8 (-23.7 ± 0.1 ‰)

236

and for (-)α-HCH were T8 (-22.4 ± 0.3 ‰), Q1 (-23.3 ± 0.3 ‰), T3 (-24.3 ± 0.3 ‰). These

237

sampling wells are all located in the area of the regional subglacial channel suggesting that more

238

degradation took place in this area (Figure 1).

239

Enantiomer fractionation: Enantiomeric fractionation of α-HCH can serve as an indicator for

240

biodegradation

241

systems, (-)α-HCH was preferentially degraded with the EF (+) value >0.5. In the Quaternary

242

aquifer, the EF (+) value ranged from 0.5 to 0.83 and the Tertiary aquifer from 0.5 to 0.85.

243

Compared to the muck samples, both aquifers have a significant shift of EF value up to 0.35

244

(well T1) which suggested biodegradation of α-HCH (Figure 3) resulting in a relative enrichment

245

of the (+)α-HCH in the residual fraction. A more intense enantiomeric fractionation was found in

246

the area of the subglacial channel compared to the area close to the dumpsite (SI-Table S2).

247

The differences in enantiomer composition and the enrichment of carbon isotopes of the

248

individual enantiomers of α-HCH suggest that biological transformation processes were active in

249

the aquifer system. Both enantiomer and isotope fractionations show similar trends indicating

250

biological processes.

26, 29, 34

. Enantiomer enrichment was observed in both aquifers. In both aquifer


Page 13 of 27


251

Evaluation of in situ transformation by combining carbon stable isotope and enantiomer

252

fractionations:

253

We established a model for the combined interpretation of the observed isotope and enantiomer

254

fractionation patterns (see SI S7). EF and isotope fractionation patterns observed in laboratory

255

studies 35, 42 provided a conceptual scheme for interpreting the degradation processes of HCH at a

256

field site with a complex history with respect to contamination and fluctuating geochemical

257

conditions (Fig. 4).

258

Figure 4 (p26)

259

We correlated the enantiomeric and isotope fractionation using fractionation pattern of

260

Sphingobium japonicum strain UT26 and Sphingobium indicum strain B90A 35. The curves show

261

already that isotope fractionation and EF are not linearly correlated for aerobic degradation.

262

Anaerobic degradation of Clostridium pasteurianum DSM525

263

enantiomeric fractionation and the fractionation factor used here needs to be taken with caution

264

as the uncertainty in the reference experiment was high and only one fractionation factor is

265

available thus far (Fig. 4). The chemical reactions studied previously were not associated with

266

enantiomeric fractionation 34. The correlation of EF and isotope fractionation of enantiomers was

267

used diagnostically to characterize degradation processes.

268

From Figure 4, wells such as T2 and T8 are close to the area of aerobic degradation which spans

269

between the fractionation pattern of Sphingobium japonicum strain UT26 and Sphingobium

270

indicum strain B90A 35. In contrast, fractionation patterns found in T4 and Q7(2014), are closer

271

to the area of anaerobic degradation of Clostridium pasteurianum DSM525

272

reaction 34. Compared to the current geochemical conditions, the model fits well for these wells.

273

However, there are also several data points which did not fit. For example, well T1 today is

33


was not associated with strong

33

or chemical


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anoxic but our model indicates aerobic degradation. Concerning the fluctuating geochemical

275

conditions, this may be because the main degradation had taken place under oxic conditions in

276

the period of lignite mining where the water table was much lower than today. Similarly well T3

277

is in the anaerobic area of our plot but the geochemical data suggested oxic condition. This

278

interpretation may also be due to the limited information on the enantiomeric and isotope

279

fractionation of α-HCH during biodegradation currently available. For an improved analysis of

280

degradation processes in the field, more reference fractionation experiments, particularly with

281

anaerobic microbial cultures, are needed.

282

Implication for the combination of CSIA and EF for the evaluation of contaminated field

283

sites:

284

α-HCH enantiomer fractionation in sewage sludge during anaerobic degradation

285

groundwater 28 with different enantiomer selectivity were reported previously. To the best of our

286

knowledge, this is the first report on the combination of enantiomer fractionation and isotope

287

fractionation of α-HCH at a field site. In our study, preferential degradation of (-)α-HCH was

288

observed which is in agreement with the previous study which applied enantiomer fractionation

289

to characterize the degradation in groundwater 28. Here, the preferential degradation of (-)α-HCH

290

was correlated with a decrease in concentration and redox potential leading to the interpretation

291

of anaerobic degradation

292

fractionation and stable isotope fractionation for chiral pesticides degradation was proposed by

293

Jin & Rolle

294

fractionation during α-HCH biodegradation by different bacterial strains and under different

295

redox conditions 36.

36

28

26

and in

. A recent study about the joint interpretation by enantiomer

. The proposed approach was illustrated by enantiomer fractionation and isotope


Page 15 of 27


296

In this study, we investigated the natural attenuation of HCHs in the groundwater. Isotope

297

fractionation and enantiomer fractionation were used to evaluate the transformation of α-HCH

298

qualitatively. A model was set up for the evaluation of α-HCH degradation by combination of

299

CSIA and EF. From geochemical data, the pH in the aquifer was neutral to acidic, which

300

suggested that chemical degradation of HCH is unlikely

301

process for HCHs transformation. The isotope fractionation indicated overall transformation and

302

the enantiomer fractionation supported in situ biotransformation.

303

To test the potential for quantification, we calculated the extent of biodegradation from isotope

304

fractionation and EF. In the first approach, the isotope enrichment between the source (the

305

average value of three muck samples from two suspected source area) and monitoring wells was

306

used to calculate the amount of biodegradation of α-HCH employing carbon isotope

307

fractionation factors from laboratory experiments

308

are reported in SI table S3. Calculated degradation of (-)α-HCH was from 30 % (T4) to 96 %

309

(T8), and (+)α-HCH was 35 % (Q7) to 91 % (Q8).

310

In the second approach, the degradation was calculated by enantiomer fractionation using Eq. 4

311

(BEF %) and EF factors from laboratory studies

312

biodegradation percentage in the aquifer is shown in SI Table S4. A degradation of (-)α-HCH

313

was between 11 % (Q2, T4 and T5) to nearly 100 % (T1 and T2), and for (+)α-HCH from 7 %

314

(Q2, T4 and T5) to 100 % (T1 and Q1) were estimated. More examples for assessment on in-situ

315

degradation were calculated for other wells (Q2, 2014_Q7, T3, T4 see SI-table S3).

316

To the best of our knowledge, chemical degradation may cause isotope fractionation but no

317

enantiomer fractionation

34

35

35

43

and biodegradation was the main

using eq. 3 (BISO %). The calculation results

were used for calculation. The calculated

. However, under acidic and neutral conditions as found at the field 15 ACS Paragon Plus Environment


Page 16 of 27

318

site, chemical degradation such as hydrolysis can be almost ruled out 43. Overall, the calculation

319

of the extent of biodegradation by isotopes and EF do not correspond. A higher value of BISO %

320

may indicate that there is a contribution of biodegradation from processes with minor

321

enantiomeric fractionation which may be unexplored thus far. In addition this may also shows

322

the uncertainty of both concepts to quantify in situ transformation of α-HCHs at this site as both

323

calculations do not correspond.

324

In contrast, the calculation of biodegradation by BEF % gives higher values than BISO %. For

325

example in well T1, where the calculation of B ISO % was found to be in the range of 80 ~ 89 %

326

((+)α-HCH), and 66 ~ 77 % ((+)α-HCH) compared to 100 % ((-)α-HCH) and 97 ~ 100 % ((+)α-

327

HCH) by BEF % (SI Table S3 and S4). This observation indicated that the enantiomer

328

fractionation and isotope fractionation were not directly correlated. The degradation quantified

329

by enantiomeric fractionation may therefore overestimate biodegradation. This may be due to the

330

variability in the microbial communities involved in the transformation with different isotope

331

and enantiomeric fractionation. For example, degradation under different geochemical conditions

332

may lead to different enantiomer selectivity, as result from changing microbial communities

333

which may leave varying isotope and enantiomeric footprints in the residual fraction

334

complicating the quantitative assessment of biodegradation. However, both methods may be

335

used diagnostically for qualitatively analyzing in situ biodegradation mechanisms and

336

differences may illustrate the significant uncertainty associated with quantitative estimation of

337

biodegradation. However as both processes are not directly correlated they could each also been

338

taken as an individual line of evidence for transformation of α-HCH.


Page 17 of 27


339

From the results, our model (Figure 4) can be applied to identify the main α-HCH degradation

340

processes for some of the sampling wells. The changing ground water table and hydrological

341

systems have probably led to changing biogeochemical conditions with oxic and anoxic periods

342

complicating the interpretation of data. Overall, the change in enantiomer and isotope patterns

343

can be a potential method for the analysis of α-HCH degradation pathways together with

344

biogeochemical conditions in the aquifer.

345

In summary, we present a model which principally can be applied to assess pathways

346

contributing to α-HCH removal at contaminated field sites including groundwater, soil and

347

atmosphere but also in engineered systems such as bioreactors. However, there are some

348

limitations need considerations, such as whether there is significant enantiomer fractionation, the

349

history of geochemical conditions and so on. With more pathway specific isotope and

350

enantiomer fractionation factors, particularly for anoxic conditions the assessment of in situ

351

biodegradation by this model could be improved. Furthermore, this model could possibly be

352

developed for other chiral pesticides and pharmaceuticals for the analysis of degradation

353

processes in the environment.

354

AUTHOR INFORMATION

355

Corresponding Author

356

*Ivonne Nijenhuis Phone: ++49 341 2351356; Fax: ++49 341 235 450822; e-mail:

357

[email protected]



Page 18 of 27

358

ACKNOWLEDGEMENTS

359

The Landesanstalt für Altlastenfreistellung Sachsen –Anhalt (Evelyn Schaffranka) is gratefully

360

acknowledged for supporting this project. The ÖGP Bitterfleld Wolfen provided geochemical

361

parameters and concentration data of HCHs and support sampling of groundwater samples in the

362

context of their routine monitoring campaigns. The GICON GmbH is acknowledged for making

363

the data of the ÖGP Bitterfeld available for this project. The Chemiepark Bitterfeld-Wolfen

364

GmbH (Dr. Michael Polk) supported the collection of HCHs muck samples. We are grateful for

365

the fellowship of Yaqing Liu from the China Scholarship Council (File No. 201306660002) and

366

University of Agriculture, Faisalabad, Pakistan for the fellowship of Safdar Bashir.

367

Julian Renpenning, Marlen Pöritz and Oliver Thiel (sampling in 2012) are acknowledged for

368

support during sampling and preparation of campaign. Matthias Gehre, Steffen Kümmel and

369

Ursula Günther are acknowledged for continuous analytical support in the Isotope Laboratory of

370

the Department of Isotope Geochemistry.

371

REFERENCE

372 373 374 375 376 377 378 379 380 381 382 383 384

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20. Meckenstock, R. U.; Morasch, B.; Griebler, C.; Richnow, H. H., Stable isotope fractionation analysis as a tool to monitor biodegradation in contaminated acquifers. J Contam Hydrol 2004, 75, (3-4), 215-55. 21. Chartrand, M.; Passeport, E.; Rose, C.; Lacrampe-Couloume, G.; Bidleman, T. F.; Jantunen, L. M.; Sherwood Lollar, B., Compound specific isotope analysis of hexachlorocyclohexane isomers: a method for source fingerprinting and field investigation of in situ biodegradation. Rapid Commun. Mass. Sp. 2015, 29, (6), 505-514. 22. Bashir, S.; Hitzfeld, K. L.; Gehre, M.; Richnow, H. H.; Fischer, A., Evaluating degradation of hexachlorcyclohexane (HCH) isomers within a contaminated aquifer using compound-specific stable carbon isotope analysis (CSIA). Water Res. 2015, 71, 187-196. 23. Ivdra, N.; Fischer, A.; Herrero-Martín, S.; Giunta, T.; Bonifacie, M.; Richnow, H. H., Carbon, hydrogen and chlorine stable isotope fingerprinting for forensic investigation on Hexachlorocyclohexanes. Environ. Sci. Technol. Submitted. 24. Elsner, M.; Zwank, L.; Hunkeler, D.; Schwarzenbach, R. P., A new concept linking observable stable isotope fractionation to transformation pathways of organic pollutants. Environ. Sci. Technol. 2005, 39, (18), 6896-6916. 25. Nijenhuis, I.; Renpenning, J.; Kümmel, S.; Richnow, H. H.; Gehre, M., Recent advances in multi-element compound-specific stable isotope analysis of organohalides: Achievements, challenges and prospects for assessing environmental sources and transformation. Trends Environ. Anal. Chem. 2016, 11, 1-8. 26. Müller, T.; Kohler, H.-P., Chirality of pollutants—effects on metabolism and fate. Appl. Microbiol. Biotechnol. 2004, 64, (3), 300-316. 27. Buser, H.-R.; Mueller, M. D., Isomer and enantioselective degradation of hexachlorocyclohexane isomers in sewage sludge under anaerobic conditions. Environ. Sci. Technol. 1995, 29, (3), 664-672. 28. Law, S. A.; Bidleman, T. F.; Martin, M. J.; Ruby, M. V., Evidence of enantioselective degradation of α-hexachlorocyclohexane in groundwater. Environ. Sci. Technol. 2004, 38, (6), 1633-1638. 29. Suar, M.; Hauser, A.; Poiger, T.; Buser, H. R.; Muller, M. D.; Dogra, C.; Raina, V.; Holliger, C.; van der Meer, J. R.; Lal, R.; Kohler, H. P., Enantioselective transformation of alpha-hexachlorocyclohexane by the dehydrochlorinases LinA1 and LinA2 from the soil bacterium Sphingomonas paucimobilis B90A. Appl. Environ. Microb. 2005, 71, (12), 85148518. 30. Padma, T. V.; Dickhut, R. M.; Ducklow, H., Variations in α‐Hexachlorocyclohexane enantiomer ratios in relation to microbial activity in a temperate estuary. Environ. Toxicol. Chem. 2003, 22, (7), 1421-1427. 31. Ridal, J. J.; Bidleman, T. F.; Kerman, B. R.; Fox, M. E.; Strachan, W. M., Enantiomers of α-hexachlorocyclohexane as tracers of air-water gas exchange in Lake Ontario. Environ. Sci. Technol. 1997, 31, (7), 1940-1945. 32. Genualdi, S. A.; Hageman, K. J.; Ackerman, L. K.; Usenko, S.; Massey Simonich, S. L., Sources and fate of chiral organochlorine pesticides in western US national park ecosystems. Environ. Toxicol. Chem. 2011, 30, (7), 1533-1538. 33. Badea, S. L.; Vogt, C.; Gehre, M.; Fischer, A.; Danet, A. F.; Richnow, H. H., Development of an enantiomer-specific stable carbon isotope analysis (ESIA) method for assessing the fate of alpha-hexachlorocyclo-hexane in the environment. Rapid Commun. Mass. Sp. 2011, 25, (10), 1363-1372. 20 ACS Paragon Plus Environment

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34. Zhang, N.; Bashir, S.; Qin, J.; Schindelka, J.; Fischer, A.; Nijenhuis, I.; Herrmann, H.; Wick, L. Y.; Richnow, H. H., Compound specific stable isotope analysis (CSIA) to characterize transformation mechanisms of α-hexachlorocyclohexane. J. Hazard. Mater. 2014, 280, 750-757. 35. Bashir, S.; Fischer, A.; Nijenhuis, I.; Richnow, H. H., Enantioselective carbon stable isotope fractionation of hexachlorocyclohexane during aerobic biodegradation by Sphingobium spp. Environ. Sci. Technol. 2013, 47, (20), 11432-11439. 36. Jin, B.; Rolle, M., Joint interpretation of enantiomer and stable isotope fractionation for chiral pesticides degradation. Water Res. 2016, 105, 178-186. 37. Walkow, F.; Enders, K.; Peklo, P. In The pollution of soil and groundwater in Bitterfeld, SAFIRA—Abstracts of the Workshop of November, 1999; 1999; pp 17-18. 38. Wycisk, P.; Neumann, C.; Gossel, W., Flooding Induced Effects from the Mining Lake Goitzsche on Groundwater and Land-use in the Bitterfeld Area. Acta Hydroch. Hydrob. 2005, 33, (5), 507-518. 39. Coplen, T. B.; Brand, W. A.; Gehre, M.; Gröning, M.; Meijer, H. A.; Toman, B.; Verkouteren, R. M., New guidelines for δ 13C measurements. Anal. Chem. 2006, 78, (7), 24392441. 40. Harner, T.; Wiberg, K.; Norstrom, R., Enantiomer fractions are preferred to enantiomer ratios for describing chiral signatures in environmental analysis. Environ. Sci. Technol. 2000, 34, (1), 218-220. 41. Jammer, S.; Voloshenko, A.; Gelman, F.; Lev, O., Chiral and isotope analyses for assessing the degradation of organic contaminants in the environment: Rayleigh dependence. Environ. Sci. Technol. 2014, 48, (6), 3310-3318. 42. Zhang, N.; Bashir, S.; Qin, J. Y.; Schindelka, J.; Fischer, A.; Nijenhuis, I.; Herrmann, H.; Wick, L. Y.; Richnow, H. H., Compound specific stable isotope analysis (CSIA) to characterize transformation mechanisms of alpha-hexachlorocyclohexane. J. Hazard. Mater. 2014, 280, 750757. 43. Liu, X.; Peng, P. a.; Fu, J.; Huang, W., Effects of FeS on the transformation kinetics of γhexachlorocyclohexane. Environ. Sci. Technol. 2003, 37, (9), 1822-1828.



Page 22 of 27

504 505

FIGURES and TABLES

506

Table 1.The concentrations of HCHs from campaigns in 2012 and 2014. The location of the

507

wells are indicated and shown in Figure 1.

508 509

Samples* 2012_Q1 2012_Q2 2012_Q3

α-HCH (μg L-1) 3.75 242 n.d.

β-HCH (μg L-1) 1.4 44 0.02

γ-HCH (μg L-1) 0.86 236 0.05

δ-HCH (μg L-1) 5.5 264 0.89

2014_Q4 2012_Q5 2012_Q6 2012_Q7 2014_Q7 2012_Q8 2012_Q9 2012_T1 2012_T2 2012_T3 2014_T4 2014_T5 2014_T6

0.38 1.86 n.d. 5.55 6.50 1.26 0.35 2.32 1.98 19.8 15.6 0.42 0.34

n.d. 0.69 1.78 3.1 2.10 n.d. 0.51 0.22 0.13 3.4 4.5 n.d. n.d.

n.d. 0.16 < 0.02 0.51 0.60 n.d. 0.02 < 0.02 0.46 20.8 n.d. n.d. n.d.

n.d. 0.58 0.12 0.8 0.40 n.d. 0.11 2.14 0.44 31.6 n.d. n.d. n.d.

2014_T7 2012_T8

0.11 4.6

n.d. 0.11

0.13 0.12

0.24 0.11

n.d.: not detected;

Compound Specific and Enantioselective Stable Isotope Analysis as

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