# Stable carbon and hydrogen isotope analysis applied to follow the anaerobic biodegradation of trimethylbenzenes in groundwater

Ouassim Boukaroum, Carine Demelas, Laurent Vassalo, Patrick Höhener

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## To cite this version:

Ouassim Boukaroum, Carine Demelas, Laurent Vassalo, Patrick Höhener. Stable carbon and hydrogen isotope analysis applied to follow the anaerobic biodegradation of trimethylbenzenes in groundwater. Journal of Contaminant Hydrology, 2025, 274, pp.104666. 10.1016/j.jconhyd.2025.104666. hal-05144759

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Submitted on 29 Jan 2026

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# Stable carbon and hydrogen isotope analysis applied to follow the anaerobic biodegradation of trimethylbenzenes in groundwater

Ouassim Boukaroum, Carine Demelas, Laurent Vassalo, Patrick Höhener*

Aix Marseille Univ, LCE, Marseille, France

## ARTICLE INFO

**Keywords:**

Isotopes

Enrichment

CSIA

Alkylbenzenes

BTEX

Groundwater

Microcosm

## ABSTRACT

Trimethylbenzene (TMB) isomers (1,2,3-TMB, 1,2,4-TMB, 1,3,5-TMB) are used as conservative tracers of alkylbenzenes in contaminated aquifers, but their resistance to anaerobic biodegradation remains uncertain. In this study, TMB degradation was investigated using enrichment cultures in microcosms under denitrifying, iron-, manganese- and sulfate-reducing conditions. Evidence for TMB biodegradation was gathered using concentration measurements, metabolites identification and multi-elemental compound specific isotope analysis (ME-CSIA). Degradation was >80 % under all anaerobic conditions and followed first-order reactions, after fluctuating lag phases depending on the condition and compound. In a mixture, a selective degradation pattern was observed with 1,2,4-TMB degradation starting almost immediately, whereas 1,3,5-TMB and 1,2,3-TMB degradation began only after 1,2,4-TMB was fully consumed. First-order biodegradation rates ranged from 0.03 ± 0.00 to 0.16 ± 0.02 day⁻¹ for 1,2,3-TMB, from 0.16 ± 0.01 to 0.32 ± 0.05 day⁻¹ for 1,2,4-TMB and from 0.02 ± 0.00 to 0.11 ± 0.02 day⁻¹ for 1,3,5-TMB. Dimethylbenzoic acid isomers were detected, which are signature metabolic intermediates formed during the anaerobic degradation of TMBs via the fumarate addition pathway. Low carbon and hydrogen fractionation, associated with their strong variability, tends to suggest that non-fractionating steps provoke isotope fractionation masking. ME-CSIA permitted to reveal that TMB biodegradation is a complex multi-step process, with A ranging from 14 ± 9 to 107 ± 91 for 1,2,3-TMB, from 29 ± 6 to 120 ± 70 for 1,2,4-TMB and from 14 ± 6 to 58 ± 23 for 1,3,5-TMB. The results will allow to identify non-conservative behaviors of TMBs during reactive transport in groundwater.

## 1. Introduction

Massive usage of petroleum hydrocarbons led to frequent oil spills in soil and groundwater (Panagos et al., 2013). Through their water solubility, mobility and toxicity, monoaromatic hydrocarbons are a major environmental and health concern in petroleum spills. In response, many remediation techniques were developed to manage hydrocarbon pollution in aquifers (Abrahams, 2002; Fent, 2004; Ossai et al., 2020). Monitored natural attenuation (MNA) is a recognized strategy to remediate hydrocarbon pollution in contaminated sites. It refers to the significant reduction of pollutants in terms of concentration, mobility and toxicity in soil and groundwater due to natural physical, chemical and biological processes. The latter are considered as the key phenomenon responsible for the degradation of contaminants and lead in the best cases to complete mineralization into innocuous products. Before MNA implementation, strong lines of evidence must demonstrate that the technique is suitable for the considered contaminated aquifer (U.S.

EPA, 1999; National Research Council (U.S.), 2000; Wiedemeier, 2007; Bombach et al., 2010; Declercq et al., 2012).

Compound Specific Isotope Analysis (CSIA) is nowadays a well implemented technique to monitor the progress of remediation after an oil spill. It relies on changes of stable isotope ratios in the remaining fraction of pollutants resulting from various processes such as biodegradation, volatilization or diffusion. CSIA has been recommended by the US EPA to justify MNA implementation at field scale by identifying the process involved and quantifying the extent of the remediation (Hunkeler et al., 2008; Aelion et al., 2010; Elsner, 2010).

In situ biodegradation of monoaromatic hydrocarbons such as benzene, toluene, ethylbenzene and xylenes (BTEX) has been highlighted at many sites worldwide with the use of CSIA (Mancini et al., 2002; Richnow et al., 2003; Meckenstock et al., 2004; Bouchard et al., 2008). While they represent a significant fraction in crude oil, only limited data is available for other monoaromatic hydrocarbons including ethyltoluene or trimethylbenzene (TMB) isomers. The last mentioned were

<footnote>* Corresponding author.
E-mail address: patrick.hohener@univ-amu.fr (P. Höhener).</footnote>
sometimes used as conservative tracers of BTEX due to their similar Henry's Law constant, sorption coefficient and their relative resistance to anaerobic biodegradation (Wiedemeier et al., 1996; Klonowski et al., 2005). However, many studies have shown inconsistent results on the biodegradability of TMBs (Peng et al., 2021). For instance, the data collected by (Aronson and Howard, 1997) indicate that 1,3,5-TMB is recalcitrant under almost all anaerobic conditions. Moreover, (Chen et al., 2009) showed that TMB biodegradation is insignificant under nitrate and sulfate reducing conditions. On the other hand, several working groups presented degradation of 1,2,4-TMB at field scale (Barker et al., 1986; Richnow et al., 2003). Also in laboratory, Haner et al. (1997) demonstrated that 1,2,4-TMB and 1,3,5-TMB can be degraded under N₂O-reducing conditions, whereas Fichtner et al. (2019) observed a complete degradation in presence of sulfate or nitrate for all TMBs. Previous studies also revealed the biodegradation of TMBs in presence of iron (Zheng et al., 2001; Chen et al., 2009). To this day, only two studies report on strictly anaerobic biodegradation of 1,2,3-TMB (Chen et al., 2008; Fichtner et al., 2019). Those contradictory results highlight the need to conduct further investigation on biological elimination of TMBs in groundwater and the potential of C and H CSIA to prove the biodegradation.

The aim of this work is to assess the biodegradability of TMBs in groundwater under anaerobic conditions. Biodegradation experiments were conducted in microcosms reproducing a contaminated aquifer in the South of France showcased in a previous study where MNA started in 2014 (Ponsin et al., 2014). The main anaerobic oxidoreductive conditions encountered in the chosen field site were simulated in laboratory microcosms: Nitrate, manganese, iron and sulfate reduction. CSIA was employed to evaluate the occurrence of degradation and to gain insight into the underlying mechanisms through quantification of enrichments factors. This approach was complemented by quantification of reaction rates, hydrogeochemical parameters and identification of biodegradation products.

# 2. Experimental section

## 2.1. Field site description

A precise description of the site has been given by Ponsin et al. (2015). Located in the South of France, the Crau alluvial plain was a confluent of the Durance and the Rhône. The site is an alluvial fan constituted mainly of deposits coming from the Alps in the form of fractured Puddingstone. Pastoralism was for centuries the main activity in the area nowadays grazed by small sheep flocks. Thus, the aquifer is characterized by low concentration of phosphate and nitrate (Naudet et al., 2004; Ponsin et al., 2014, 2015).

In 2009, the site has been contaminated by more than 5000 m³ of crude oil following the break of an underground pipeline. The pollution reached the underlying aquifer forming an oil lens between 8 and 11 m of depth. During the first weeks of the pollution, remediation actions consisted in off-site treatment and replacement of contaminated surface soil. A hydraulic barrier was installed to confine the pollution. Also, a pump-and-skim system was installed with reinjection after activated charcoal treatment coupled with dual-phase oil extraction. Since 2014, groundwater pollution is treated by MNA mainly under sulfate-reducing conditions, although all other anaerobic conditions were also present locally (Ponsin et al., 2015). The field site has been the object of several studies, allowing for frequent monitoring and sampling of the pollution. During the second year after the spill, oil and microbial sludge were collected for a microcosm degradation study aiming at elucidating the phosphorus limitation of degradation (Ponsin et al., 2014). Oil and sludge were also stored in the dark at 4 °C until used for microbial inoculum in the present study.

## 2.2. Chemicals and reagents

2,2,4,4,6,8,8-Heptamethylnonane (>98 %), 1,2,3-TMB (>99.8 %), 1,2,4-TMB (>98 %), 1,3,5-TMB (>98 %), sodium azide (>99.5 %) were from Sigma-Aldrich. NaNO₃ (Fluka Analytical, >99 %), Na₂SO₄ (Thermo Scientific, >99 %), Na₂HPO₄, 12 H₂O (AnalàR NORMAPUR, >99 %), Limonite (FeO(OH)) and Cryptomelane (K(Mn⁴⁺, Mn²⁺)₈O₁₆) were used to supply iron and manganese oxides. Aquifer sediments were collected near the spill site of the Crau aquifer plain in an uncontaminated quarry at around 9 m of depth (Ponsin et al., 2015). Organic matter was removed by addition of hydrogen peroxide (Fisher Chemicals, 35 wt% solution in water, stabilized) mixed with drops of ethanol (Sigma Aldrich, Purity>96.0–97.2). The aquifer material was then rinsed with distilled water, dried at 110 °C overnight and thereafter fractionated by size. The fraction larger than 2 mm was removed by dry sieving. The remaining sediment was wet filtrated on a mesh of 0.124 mm, rinsed with distilled water and dried at 110 °C for at least 24 h. The aquifer sediment characteristics have been measured in a previous study and are given back in Table SI 1.

## 2.3. Biodegradation experiment

Anaerobic bacterial cultures were enriched on TMBs by incubating 1 mL of microbial sludge in 100 mL of synthetic groundwater and 10 g of aquifer sediment in 125 mL serum bottles (Supelco) sealed with PTFE Mininnert valves (Supelco) (Fig. SI 1). Synthetic groundwater mimicking the Crau aquifer was used as a growth medium allowing to precisely control the redox condition involved by adding only one electron acceptor in the water. The medium was supplemented in excess of electron acceptors to ensure that they are not a growth limiting factor. For nitrate and sulfate reducing conditions the medium was supplemented with 150 mg L⁻¹ of either NO₃⁻ (NaNO₃) or SO₄²⁻ (Na₂SO₄). Iron and manganese reducing conditions were promoted by supplying respectively 1 g of thinly crushed (< 2 mm) limonite or cryptomelane. All microcosms were amended with 10 mg L⁻¹ of phosphate (Na₂HPO₄, 12 H₂O) to shorten the lag phase (Ponsin et al., 2014). Detailed compositions of microcosms under each studied redox condition can be found in Table SI 2.

Following the enrichment procedure described by (Haner et al., 1997), 2 mL of 2,2,4,4,6,8,8-heptamethylnonane (HMN) containing 3.8 % (mol/mol) of the three TMBs isomers were added. HMN was used as an inert carrier allowing a continuous dissolution of TMBs while avoiding toxic effect (Rabus et al., 1993; Haner et al., 1997). The proportion of TMBs in HMN were calculated to be representative of the concentration found in the oil plume in the Crau aquifer. The oxygen in the headspace was removed by flushing with nitrogen gas. The microcosms were stored at in situ temperature of 18–20 °C under water in the dark to preserve the anaerobiosis and avoid the development of algae (Fig. SI 2). The bacterial activity was monitored by following the electron acceptor depletion and by comparison with abiotic controls.

After an incubation time going from 30 to 200 days, 20 mL of water containing the bacterial pre-cultures were transferred into fresh microcosms containing 80 mL of synthetic groundwater and no HMN. The medium was spiked with a mixture of 20 mg L⁻¹ of each TMB isomer. Each redox condition included 3 active batches and 3 control microcosms. The latter did not receive bacterial precultures and contained thus 100 mL of synthetic groundwater spiked with 250 mg L⁻¹ of sodium azide to inhibit bacterial growth.

Hydrogeochemical characterization was performed at the start and end of the biodegradation. Details about major anions and cations, pH and alkalinity measurements are given in SI part 1.6. Anaerobic biodegradation by-products were analyzed by GC-MS after liquid-liquid extraction followed by a derivatization reaction with N,O-bis-(trimethylsilyl)trifluoroacetamide (BSTFA) at the end of the biodegradation. Details about the extraction, derivatization and GC-MS procedures can be found in SI part 1.7.
## 2.4. Concentration measurement

For TMBs concentration, 1 mL of water was collected through the Mininert valve with sterilized syringes and introduced in 20 mL headspace vial (Sigma Aldrich) filled with NaCl saturated solution for sample stabilization and headspace extraction. Samples were spiked with toluene, which serves as an internal standard, and tightly sealed using aluminum crimp caps with PTFE/Silicone liner (Supelco). Samples were stored at -8 °C if not analyzed within 24 h.

TMBs were analyzed by GC-MS using a Clarus 580 GC connected to a Clarus SQ 8 T mass spectrometer after a headspace extraction made with a TurboMatrix headspace sampler HS 40, all from Perkin Elmer Inc. (USA). Details about the HS-GC-MS method are provided in SI part 1.4. The average measured concentrations and stable isotopes for each compound were corrected with respect to concentrations in abiotic microcosms. The bias from mass removal by repetitive sampling was corrected using the stepwise correction method (Buchner et al., 2017).

## 2.5. Compound specific isotope analysis

To measure stable isotope ratios in TMBs, 1 mL of water was collected for carbon analysis and 2 mL for hydrogen analysis. Sample collection was performed using sterilized syringes through the Mininert valve of microcosms. For carbon and hydrogen isotope measurements, samples were respectively introduced into 10 mL and 20 mL headspace vial (Thermo Fisher Scientific), then filled up to 6 mL and 15 mL with NaCl saturated solution for sample stabilization and headspace extraction. Samples were sealed using magnetic screw caps with Silicone/ PTFE (Supelco) and stored upside down at -8 °C if not analyzed within 24 h.

Headspace extraction was performed using a Triplus RSH autosampler (Thermo Fisher Scientific). The samples were heated at 80 °C for 10 min while shaking at 250 rpm using an HS oven for six HS vials (Thermo Fisher Scientific). For carbon measurement, the autosampler was equipped with a gas-tight HS syringe (100-2500 µL) without closing valve. After equilibration, 1 mL of headspace was collected by the heated syringe and injected into the GC. For hydrogen measurement, headspace Solid Phase Micro Extraction (SPME) injection was applied due to higher detection limits of the IRMS system. A Carbon Wide Range/PDMS fiber (10 mm × 95 µm) was exposed 10 min into the headspace above the sample. Analytes were then desorbed directly into the GC injector heated at 280 °C.

Compound specific isotope ratios were measured using a Delta V Advantage Isotope Ratio Mass Spectrometer connected via an Isolink II interface to a Trace 1310 GC, all from Thermo Fisher Scientific. Chromatographic separation was performed with a TG-5MS GC column (30 m × 0.25 mm, 0.25 µm film thickness). Samples were injected via an iConnect PTV (Thermo Scientific) injector at 280 °C in splitless mode with helium as carrier gas at 1.5 mL.min⁻¹ constant flow rate. The GC oven was programmed in ramp mode allowing for TMBs separation: starting at 40 °C maintained 3 min, heating at 50 °C.min⁻¹ to 120 °C maintained 12 min, followed by another 50 °C.min⁻¹ ramp to 290 °C maintained 7 min.

For carbon isotope measurement, the oxidation furnace constituted of Pt, Ni and CuO was set to 1000 °C. The furnace was oxidized before each experiment with a stream of O₂ for 120 min. For hydrogen measurement, a second high temperature conversion reactor was set to 1400 °C. VOC concentrations were recorded using the amplitude of mass 44 of CO₂ in IRMS in combustion mode or the amplitude of mass 2 of H₂ in pyrolysis mode.

The measured delta δ¹³C or δ²H (‰) was calculated following Eq. (1):

$$ \delta (‰) = \frac{R_{\text{sample}} - R_{\text{standard}}}{R_{\text{standard}}} \times 1000 \qquad (1) $$

With Rsample being the ratio of the sample and Rstandard the abundance

ratios of the international standards (VPDB, 0.0112372 or VSMOW, 1.5575 × 10⁻⁴ Isotope ratios of TMB standards were measured every 12 samples analyzed to characterize the daily instrumental drift.

Isotope fractionation during biodegradation of TMBs was calculated with the Rayleigh equation (Rayleigh, 1896; Aelion et al., 2010):

$$ 1000 \times \ln \left( \frac{\delta (‰) + 1000}{\delta_0 (‰) + 1000} \right) = \varepsilon_{\text{bulk}} \times \ln f \qquad (2) $$

where f is the remaining fraction of the substrate, δ₀ is the initial isotope deviation of the substrate, εbulk is the enrichment factor of the substrate.

Hydrogen and carbon isotopes were correlated in a dual element isotope plot, characterized by the $\Lambda^{\text{H}}$ value:

$$ \Lambda^{\text{H}} = \frac{\Delta\delta^2 H}{\Delta\delta^{13}C} \approx \frac{\varepsilon_{\text{H,bulk}}}{\varepsilon_{\text{C,bulk}}} \qquad (3) $$

where Δδ is the change of isotope ratios from initial values, and εbulk are the bulk enrichment factors for hydrogen and carbon.

The method detection limits (MDLs) for δ²H and δ¹³C were calculated by the moving mean approach as described by Jochmann et al. (2006). The MDLs for each analyte is the lowest concentration within the iterative interval and for which the standard deviation after triplate measurement is lower than ±5 ‰ for δ²H and ±0.5 ‰ for δ¹³C.

To avoid bias in the experimental determination of TMB isotope enrichment factors, the stepwise correction (SW) method proposed by Buchner et al. (2017) was applied. The SW method accounts for gas-water partitioning and sample removal, which affect greatly the concentration of volatile compounds such as TMBs. The detailed correction procedure is presented in SI part 1.5. Finally, measured concentrations in active batches were corrected by the average relative concentration in abiotic controls.

## 2.6. Calculation of apparent kinetic isotope effects

To identify and confirm the trimethylbenzene anaerobic degradation pathway, the procedure described by Elsner et al. (2005) is applied. The variation of the reaction rate between stable isotopes, referred as the kinetic isotope effect (KIE) is referred as:

$$ \text{KIE}_E = \frac{h_k}{l_k} \qquad (4) $$

where h_k and l_k are respectively the pseudo-first-order rate constants of the light and heavy stable isotopes of element E. Theoretically, broken bonds produce the same KIE value independent of the molecule or the reaction. Since KIE values are not available in the literature for TMBs, comparison of the isotope fractionation can be performed with molecules of different sizes following a similar degradation pathway such as toluene. To circumvent the isotope dilution effect by nonreactive positions and the loss of the position specific KIE during the conversion step prior to IRMS measurement, the Apparent Kinetic Isotope Effect (AKIE) is calculated, enabling direct comparison with isotope effect data available in literature.

$$ \text{AKIE}_E = \frac{1}{1 + z \times \frac{n}{x} \times \frac{\varepsilon_{\text{bulk}}(‰)}{1000}} \qquad (5) $$

with n the number of atoms of element, x the number of atoms at reactive sites and z the number of indistinguishable reactive sites.

To perform comparison between $\Lambda^{\text{H}}$ with literature data, the $\Lambda^{\text{H}}_{\text{rp}}$ was computed using enrichment factors at reactive position:
$$ \Lambda_{rp}^{\frac{H}{C}} = \frac{\epsilon_{H,rp}}{\epsilon_{C,rp}} = \frac{\left(\frac{n_H}{x_H}\right) \times \epsilon_{H,bulk}}{\left(\frac{n_C}{x_C}\right) \times \epsilon_{C,bulk}} \quad (6) $$

with $\epsilon_{H,rp}$ and $\epsilon_{C,rp}$ corresponding to enrichment factors at reactive position for hydrogen and carbon, respectively.

### 3. Results and discussion

#### 3.1. Anaerobic degradation kinetics of trimethylbenzene isomers

Detailed results of terminal electron acceptors evolution during the bacterial enrichment step are presented in Fig. SI 3. Enrichment cultures were able to utilize TMBs as substrates under each anaerobic condition studied and were transferred into active microcosm batches. In active batches, the degradation of TMBs was >80% under all anaerobic conditions tested and followed a first-order reaction after initial lag phases. In abiotic controls, a maximum concentration decrease of 20% was observed (data not shown). The lag phases varied significantly depending on the condition and compound. These variations can be mainly attributed to the differing bacterial concentrations introduced in each replicate after the enrichment step. TMB degradation followed a selective degradation pattern, with 1,2,4-TMB degradation starting almost immediately, whereas 1,3,5-TMB and 1,2,3-TMB degradation began only after 1,2,4-TMB was fully consumed (Fig. 1). This selective degradation pattern has been previously reported in studies of TMBs biodegradation in presence of BTEX. Haner et al. (1997) observed that 1,2,4-TMB, 1,3,5-TMB, p-xylene and m-xylene were completely degraded under denitrifying conditions, whereas benzene, o-xylene and 1,2,3-TMB were not metabolized. Chen et al. (2009) demonstrated that under nitrate-, iron- and sulfate-reducing conditions TMB degradation

started only after BTEX concentrations were significantly reduced. When exposed to a mixture of organic compounds, bacteria typically prioritize degrading the compound with the highest biodegradability. Indeed, Meckenstock et al. (2004) demonstrated that the degradation of o-xylene was inhibited even by very low concentration of toluene. The biodegradability of alkylbenzenes strongly depends on the position of alkylation, with adjacent methyl groups reducing their susceptibility to biodegradation as highlighted by George et al. (2002). Thus, the selective degradation pattern is attributed to steric hindrance from closely positioned methyl groups, which restricts the ability of benzylsuccinate synthase (BSS) to abstract the hydrogen atom from the targeted methyl group. Due to its accessible methyl group, bacteria prioritize degrading 1,2,4-TMB before the other isomers, with no preference between 1,3,5-TMB and 1,2,3-TMB. Dou et al. (2008) observed similar behavior for xylene isomers in a high-concentration BTEX mixture, with m-xylene degrading first, followed by o-xylene and then p-xylene.

To calculate degradation rates, lag phases and endpoints were excluded, focusing solely on the linear phase of concentration decrease (Fig. SI 4). All anaerobic conditions considered, 1,2,4-TMB degradation was the fastest, while no significant difference was observed between 1,3,5-TMB and 1,2,3-TMB (Table 1). After the lag phases, TMBs were completely removed within 2 to 7 days, except for 1,3,5-TMB and 1,2,3-TMB, whose degradation was slower under iron- and manganese-reducing conditions due to the slow dissolution rates of limonite and cryptomelane minerals. Under denitrifying conditions, the calculated degradation rates were of the same order of magnitude as measured by Hutchins et al. (1991) for 1,2,4-TMB. Under sulfate-reducing conditions, 1,2,4-TMB degradation was 10 times faster than Zheng et al. (2001), which can be explained by the lower temperature. In microcosms, the degradation rate of 1,3,5-TMB is 4 to 10 times faster than that observed at the field site by Thierrin et al. (1995), which is expected since the molecule is not mixed with other compounds apart from TMBs. Under

<figure><img src="image_5.png" /><figcaption>Fig. 1. Evolution of trimethylbenzene isomers relative concentration under anaerobic conditions in microcosms. Replicates are plotted separately to showcase the significant differences in lag phases. Relative concentrations are plotted after correction by the average relative concentration in abiotic controls.</figcaption></figure>
<table><caption>**Table 1**
Summary of experimentally determined and literature-reported TMB degradation rates.</caption><thead><tr><th>Compound</th><th>Condition</th><th>Study type</th><th>pH</th><th>Temperature<br>°C</th><th>Lag phase<br>Day</th><th>Degradation rate<br>Day<sup>-1</sup></th><th>Reference</th></tr></thead><tbody><tr><td rowspan="12">1,2,4-TMB</td><td rowspan="6">Aerobic<br>Nitrate</td><td>Column</td><td></td><td>23 ± 2</td><td></td><td>4.98</td><td rowspan="12">Höhener et al. (2003)</td></tr><tr><td>Bottle</td><td></td><td>18-20</td><td>4 ± 0</td><td>0.16 ± 0.01</td></tr><tr><td>Bottle</td><td>7-8</td><td>12</td><td></td><td>0.148</td></tr><tr><td>Bottle</td><td></td><td>12</td><td></td><td>0.27</td></tr><tr><td>Bottle</td><td></td><td>12</td><td></td><td>0.00315</td></tr><tr><td>Bottle</td><td></td><td>12</td><td></td><td>0.0422</td></tr><tr><td rowspan="6">Iron<br>Manganese<br>Sulfate</td><td>Column</td><td></td><td>20</td><td></td><td>0.08</td></tr><tr><td>Column</td><td></td><td>10-12</td><td></td><td>0.19</td></tr><tr><td>Bottle</td><td>7-8</td><td>18-20</td><td>5 ± 0</td><td>0.13 ± 0.01</td></tr><tr><td>Bottle</td><td>7</td><td>23</td><td></td><td>0.006</td></tr><tr><td>Bottle</td><td>7-8</td><td>18-20</td><td>2 ± 0</td><td>0.32 ± 0.05</td></tr><tr><td>Bottle</td><td>7-8</td><td>18-20</td><td>4 ± 0</td><td>0.17 ± 0.01</td></tr><tr><td rowspan="6">Anaerobic mixed conditions<br>Nitrate</td><td>Bottle</td><td></td><td>7-8</td><td></td><td>0.009</td></tr><tr><td>Bottle</td><td></td><td>7</td><td></td><td>0.02</td></tr><tr><td>Bottle</td><td></td><td>7</td><td></td><td>0.07</td></tr><tr><td>Column</td><td></td><td>20</td><td></td><td>0.189</td></tr><tr><td>Field site</td><td>6.5-7.1</td><td>17-29</td><td></td><td>0.1 ± 0.01</td></tr><tr><td>Column</td><td></td><td>20</td><td></td><td>0.2</td></tr><tr><td rowspan="12">1,3,5-TMB</td><td rowspan="6">Nitrate</td><td>Bottle</td><td></td><td>7-8</td><td></td><td>0.21</td></tr><tr><td>Bottle</td><td></td><td>7-8</td><td></td><td>0.04 ± 0.01</td></tr><tr><td>Bottle</td><td></td><td>7</td><td></td><td>0.003</td></tr><tr><td>Bottle</td><td></td><td>7-8</td><td></td><td>0.02 ± 0.00</td></tr><tr><td>Bottle</td><td></td><td>7-8</td><td></td><td>0.11 ± 0.02</td></tr><tr><td>Bottle</td><td></td><td>7</td><td></td><td>0.05</td></tr><tr><td rowspan="6">Iron<br>Manganese<br>Sulfate</td><td>Column</td><td></td><td>20</td><td></td><td>0.0039</td></tr><tr><td>Field site</td><td></td><td>7-8</td><td></td><td>0.16 ± 0.02</td></tr><tr><td>Bottle</td><td></td><td>7-8</td><td></td><td>0.04</td></tr><tr><td>Bottle</td><td></td><td>7</td><td></td><td>0.03</td></tr><tr><td>Bottle</td><td></td><td>7-8</td><td></td><td>0.04 ± 0.01</td></tr><tr><td>Bottle</td><td></td><td>7</td><td></td><td>0.013</td></tr><tr><td rowspan="6">Nitrate</td><td>Column</td><td></td><td>20</td><td></td><td>0.04</td></tr><tr><td>Column</td><td></td><td>10-12</td><td></td><td>0.03</td></tr><tr><td>Bottle</td><td></td><td>7-8</td><td></td><td>0.04 ± 0.01</td></tr><tr><td>Bottle</td><td></td><td>7</td><td></td><td>0.013</td></tr><tr><td>Bottle</td><td></td><td>7-8</td><td></td><td>0.03 ± 0.00</td></tr><tr><td>Bottle</td><td></td><td>7-8</td><td></td><td>0.15 ± 0.02</td></tr><tr><td rowspan="6">Nitrate</td><td>Bottle</td><td></td><td>7-8</td><td></td><td>0.11</td></tr><tr><td>Bottle</td><td></td><td>7-8</td><td></td><td>0.11</td></tr><tr><td>Bottle</td><td></td><td>7</td><td></td><td>0.11</td></tr><tr><td>Bottle</td><td></td><td>7</td><td></td><td>0.11</td></tr><tr><td>Bottle</td><td></td><td>7</td><td></td><td>0.11</td></tr><tr><td>Bottle</td><td></td><td>7</td><td></td><td>0.11</td></tr></tbody></table>

<figure><img src="image_6.png" /><figcaption>**Fig. 2.** A – GC-MS chromatograms of derivatized microcosm extracts in denitrifying conditions. Only the portion of the chromatogram showing the degradation products is presented. Degradation of trimethylbenzene was performed in mixture (All TMBs), and for each trimethylbenzene isomer taken separately. B – GC-MS mass spectrums of identified trimethylsilyl-derivatized dimethylbenzoic acid isomer. Colors match the GC-MS chromatogram.</figcaption></figure>
iron-reducing conditions, TMB degradation rates were 4 to 20 times higher than those reported by Chen et al. (2009). In their study, groundwater was amended with gasoline, introducing more readily biodegradable compounds such as BTEX, which likely inhibited or delayed TMB degradation due to preferential substrate utilization. Overall, the anaerobic degradation rates of TMBs remain lower than those observed under aerobic conditions, as demonstrated by Höhener et al. (2003) for 1,2,4-TMB in a column experiment. Despite negative conclusions on TMBs degradation from other laboratory studies, particularly for 1,2,3-TMB, the biodegradation of TMBs under the conditions of the Crau aquifer has proven to be highly efficient. This efficiency can likely be attributed to the relatively higher temperatures in this aquifer. According to Fichtner et al. (2019), while lower temperatures did not significantly affect the degradation of 1,2,4-TMB and 1,3,5-TMB with nitrate, a marked decline in 1,2,3-TMB degradation was observed, likely due to reduced biological activity. Unfortunately, the bacterial strains developed during this study could not have been precisely identified. In the Crau aquifer, it has been proven that the majority of the cultivable microorganisms were sulfate-reducing strains (Ponsin et al., 2014). During the enrichment phase, bacteria communities may have adapted to specifically degraded TMBs as they were the sole source of carbon available, potentially explaining their ability to degrade even the recalcitrant 1,2,3-TMB. TMB degradation could also have been the result of cometabolism. However, the strains could have lost their ability to degrade other monoaromatics hydrocarbons as a consequence of their adaptation to TMBs.

## 3.2. Identification of the anaerobic degradation pathway of trimethylbenzenes

TMB degradation products were analyzed at the end of the experiment under denitrifying conditions (Fig. 2). GC-MS analysis of the

extracts revealed three dimethylbenzoic acid (DMBA) isomers, which are signature metabolites of anaerobic degradation of alkylbenzenes. Indeed, benzoic acid methylated homologs do not naturally occur in uncontaminated aquifers and are not constituents of crude oil (Martus and Schaal, 2010). Instead, they are formed as metabolic intermediates during the degradation of alkylbenzenes, which follows the fumarate addition pathway under denitrifying, sulfate-reducing, and iron-reducing conditions (Spormann and Widdel, 2000; Ladino-Orjuela et al., 2016). Cozzarelli et al. (1990) observed the formation of similar organic acids as primary products resulting from the transformation of alkylbenzenes in an anoxic groundwater environment. Moreover, (Beller, 2000) detected the formation of methyl benzoic acid, benzylsuccinate and their methyl homologs after alkylbenzenes anerobic degradation. The DMBA isomers were also found at field scale in iron-reducing condition (Kampbell et al., 1996), sulfate-reducing condition and methanogenesis (Martus and Schaal, 2010), confirming that fumarate addition occurs regardless of the redox condition.

To accurately identify the corresponding metabolite produced, the degradation of individual TMB isomer was also conducted under denitrifying conditions. According to (Martus and Schaal, 2010), the number of possible DMBA isomers derived from TMB precursors depends on the positioning of the methyl groups on the aromatic ring. This ranges from a single structurally corresponding DMBA for 1,3,5-TMB to three distinct DMBAs for 1,2,4-TMB. However, 3,4-DMBA was the only 1,2,4-TMB metabolite detected and was present in the highest concentration among the other metabolites (Fig. 2). 3,5-DMBA, whose precursor is 1,3,5-TMB, was detected and possibly coeluted with 1,2,3-TMB metabolites, 2,3-DMBA and 2,5-DMBA, at retention time 39.17 min when the degradation is performed in mixture.

The presence of those signature metabolites allows us to identify confidently the anaerobic degradation pathway of the TMB. A simplified degradation mechanism is proposed for TMB isomers and is presented in

<figure><img src="image_7.png" /><figcaption>**Fig. 3.** Reaction mechanism for fumarate addition on 1,3,5-Trimethylbenzene catalyzed by benzylsuccinate synthase (E-Gly) according to (Morasch et al., 2004; Heider and Schühle, 2013; Ladino-Orjuela et al., 2016) and identification of reactive sites for the calculation of the apparent kinetic isotope effect (AKIE) from measured fractionation data. The proposed simplified anaerobic degradation pathway following the fumarate addition is adapted from (Harwood et al., 1998; Carmona et al., 2009). The same mechanism applies to other alkylbenzenes, with E corresponding to polypeptide chain of the enzyme.</figcaption></figure>
**Fig. 3.** Alkylbenzenes are degraded in anerobic conditions via fumarate addition, which is a stereospecific and a rate-limiting step provoking isotope fractionation as demonstrated for toluene (Himo, 2002; Morasch et al., 2002). The reaction is catalyzed by the benzylsuccinate synthase, a highly oxygen sensitive enzyme responsible for the abstraction of a hydrogen atom from a methyl group via its glycyl radical (Beller and Spormann, 1997; Krieger et al., 2001). The enzyme-bond dimethyl benzyl radical reacts with the double bond of the fumarate cosubstrate within the enzyme's active site, forming a benzylsuccinyl radical. This radical will stabilize itself by reabstracting a hydrogen from the enzyme, producing a (R)-dimethylbenzylsuccinate (Himo, 2002; Morasch et al., 2004; Heider, 2007). This intermediate can be further oxidized to benzoyl-CoA, a central aromatic intermediate in anaerobic oxidation of aromatic compounds, which undergoes ring reduction, cleavage and oxidation to CO₂ (Harwood et al., 1998; Foght, 2008).

### 3.3. Elucidation of trimethylbenzenes reaction pathway using compound specific isotope analysis

Anaerobic degradation of TMBs resulted in minimal carbon fractionation but significant hydrogen fractionation. Enrichment factors per compound and condition were computed using the Rayleigh equation (Fig. 4). Concentration and carbon isotope ratio were highly correlated ($R^2$ between 0.77 and 0.93) for 1,2,4-TMB in sulfate-reducing condition, 1,3,5-TMB under denitrifying condition and all conditions except denitrifying condition for 1,2,3-TMB. For 1,3,5-TMB, concentration and carbon isotope ratio were moderately correlated in iron-, manganese- and sulfate-reducing conditions. In contrast, 1,2,4-TMB showed poor correlation across all conditions except sulfate-reducing conditions. Although the same initial reaction occurs, fumarate addition yielded differing isotope carbon fractionation, more pronounced under sulfate-reducing conditions, with $\epsilon_{C,bulk}$ ranging from -0.6 ± 0.1 ‰ to -1.1 ± 0.1 ‰. On the other hand, iron- and manganese-reducing conditions

result in minimal or no significant carbon fractionation, with $\epsilon_{C,bulk}$ between -0.2 ± 0.1 to -0.6 ± 0.1 ‰. Interestingly, significant fractionation of -1.4 ± 0.2 ‰ under nitrate-reducing conditions has only been observed for 1,3,5-TMB. To our knowledge, only Richnow et al. (2003) observed an isotope enrichment for TMBs during anaerobic biodegradation at a field site. For 1,2,4-TMB, a carbon enrichment of -5.78 ‰ was observed along the flow path, while 1,3,5-TMB did not get enriched. This value is significantly higher than the results obtained in our study, in which carbon fractionation did not exceed 1 ‰ at most.

Concentration and hydrogen isotope fractionation were highly correlated ($R^2$ between 0.71 and 0.97) in every reducing condition tested for 1,2,4-TMB and 1,3,5-TMB, and poorly correlated across all conditions except iron-reducing condition for 1,2,3-TMB. Overall, hydrogen fractionation was significantly greater than carbon fractionation (Fig. 4). This difference arises from the greater relative mass difference between heavy and light hydrogen isotopes compared to carbon isotopes (Elsner et al., 2005). Among the three isomers, 1,2,3-TMB degradation exhibited the smallest hydrogen fractionation, with $\epsilon_{H,bulk}$ ranging from -7 ± 3 ‰ to -15 ± 10 ‰, whereas 1,3,5-TMB degradation resulted in the most pronounced hydrogen fractionation, with $\epsilon_{H,bulk}$ ranging from -11 ± 3 ‰ to -32 ± 5 ‰. In contrast to carbon isotopes, TMB anaerobic biodegradation resulted in a strong hydrogen fractionation under iron- and manganese-reducing conditions. Moreover, anaerobic degradation produced nearly identical hydrogen fractionation for each compound individually, confirming that fumarate addition occurs consistently under all anaerobic conditions for alkylbenzenes. Despite the distinct differences in the biodegradability of TMBs, no consistent trend is evident in carbon fractionation. Since isotope fractionation can only arise from isotopes located in the methyl group targeted by the BSS, the lower carbon fractionation can be attributed to an uneven isotope distribution, with heavy carbon isotope predominantly located at non-reacting positions. Indeed, Julien et al. (2016) using position specific isotope analysis on toluene, observed that

<figure><img src="image_8.png" /><figcaption>**Fig. 4.** Rayleigh plots for carbon and hydrogen isotope fractionation for trimethylbenzenes isomers under anaerobic conditions. Dashed lines represent ±0.5 ‰ and ± 5.0 ‰ significance thresholds respectively in carbon and hydrogen.</figcaption></figure>
the C-1 (i.e. the carbon directly linked to the methyl group) is more enriched than the methyl group. By analogy, TMBs isotope distribution may be uneven depending on the manufacturing process, with the aromatic ring likely more enriched than the methyl groups. We can postulate that TMBs undergo a more important fractionation in hydrogen due to the higher probability of having a deuterium located in a reactive position.

AKIE values were computed using Eq. 5 assuming TMB degradation via the fumarate addition presented in Fig. 5. Fumarate addition involves the breaking of a C—H bond during the hydrogen abstraction of one of the methyl groups by the benzylsuccinate synthase. Given that intermolecular competition occurs between the three methyl groups potentially involved in the fumarate addition, for carbon isotopes *n* = 9, *x* = 3 and *z* = 3 and for hydrogen isotopes *n* = 12, *x* = 9 and *z* = 9 (Fig. 3). Computed AKIE values were compared to values computed for fumarate addition for toluene and xylenes available in the literature (Fig. 5 and Table 2). AKIE values for carbon and hydrogen fell on the lower range of the region for fumarate addition, making this scenario consistent for TMBs anaerobic biodegradation.

The AKIE<sub>C</sub> and AKIE<sub>H</sub> for TMBs fumarate addition were overall in the lower range of observed values in literature. AKIE did not differ significantly between anaerobic conditions or TMBs (Fig. SI 9 and 10). Denitrifying conditions exhibited strong AKIE<sub>C</sub> variations between the three TMBs, ranging from 1.001 ± 0.001 ‰ for 1,2,3-TMB to 1.013 ± 0.002 ‰ for 1,3,5-TMB. Those AKIE<sub>C</sub> were drastically lower than values recorded by Vogt et al. (2008) for toluene, ranging from 1.019 ± 0.001 to 1.042 ± 0.002 ‰, or by Herrmann et al. (2009) for m-xylene, where it reached 1.019 ± 0.001 ‰. Only AKIE<sub>C</sub> for fumarate addition in sulfate-reducing conditions aligns with values reported in the literature, ranging from 1.006 ± 0.001 ‰ for 1,2,4-TMB to 1.010 ± 0.001 ‰ for 1,2,3-TMB. The lowest value observed was 1.006 ± 0.001 for o-Xylene degradation by a bacteria consortium, as reported by Herrmann et al.

(2009). Surprisingly, TMBs AKIE<sub>H</sub> values were drastically lower than those reported in the literature for fumarate addition of alkylbenzenes, except under manganese-reducing conditions. In this case, AKIE<sub>H</sub> ranged from 1.092 ± 0.043 ‰ for 1,2,3-TMB to 1.295 ± 0.040 ‰ for 1,2,4-TMB, which goes in accordance with values reported for toluene by Dorer et al. (2016), which ranged from 1.157 ± 0.032 ‰ to 1.404 ± 0.110 ‰.

AKIE for TMBs for carbon and hydrogen are also lower than the semiclassical Streitwieser limits for the breaking of a C—H bond, which are set at 1.021 for carbon and 6.4 for hydrogen (Huskey, 1991). This tends to suggest that the breaking of the C—H during the hydrogen abstraction by the BSS is not the rate-limiting step during TMBs degradation. Indeed, considering that TMB anaerobic biodegradation via the fumarate addition pathway is a multi-step process, the presence of slow non-fractionating steps makes AKIE and expected KIE differ significantly (Elsner et al., 2005). Those non-fractionating steps could explain the lower enrichment factors and AKIE observed for TMBs anaerobic biodegradation compared to literature values. Moreover, the variability of the isotope fractionation could be attributed to different non-fractionating steps between anaerobic conditions. A key distinction of our study is the use of a TMB isomers mixture, unlike previous research that focused on single-compound degradation in pure phase. Consequently, a small organic phase formed at the water surface, requiring TMBs to dissolve into water. This dissolution step can act as a bottleneck for anaerobic biodegradation and has been shown to significantly reduce isotope fractionation (Elsner, 2010). Isotope fractionation masking due to mass transfer into microbial cells can also explain why our value can be lower than theoretical values. Indeed, Tobler et al. (2008) observed smaller toluene isotope fractionation with solid iron, where microorganisms attach, compared to dissolved iron. Similarly, Dorer et al. (2016) observed a comparable effect during toluene biodegradation under manganese-reducing conditions using MnO<sub>2</sub> particles, which

<figure><img src="image_9.png" /><figcaption>**Fig. 5.** Comparison of carbon and hydrogen AKIE for fumarate addition on alkylbenzenes with available data. Dashed lines correspond to Streitwieser Semiclassical limits for isotope effects at 25 °C according to Elsner et al. (2005).</figcaption></figure>
<caption>**Table 2**</caption>

Bulk enrichment factors ($\epsilon_{bulk}$), apparent kinetic isotope effects (AKIE) and bulk lambda ($\Lambda^{H/C}$) values for anaerobic degradation of trimethylbenzenes and others alkylbenzenes following fumarate addition to a methyl group with terminal electron acceptors used in this study. Error values for AKIE and $\Lambda^{H/C}$ were calculated using error propagation from $\epsilon_{bulk}$ error. $\epsilon_{bulk}$ error in our work is computed via linear regression analysis (SI eq. 7–13). AKIE values for fumarate addition were recalculated using $\epsilon_{bulk}$ provided in the literature.

<table><thead><tr><th>Compound</th><th>Conditions</th><th>Strain</th><th>ε<sub>C,bulk</sub><br>%<sub>0</sub></th><th>AKIE<sub>C</sub><br>%<sub>0</sub></th><th>ε<sub>H,bulk</sub><br>%<sub>0</sub></th><th>AKIE<sub>H</sub><br>%<sub>0</sub></th><th>Λ<sup>H/C</sup><br>%<sub>0</sub></th><th>Λ<sup>H/C</sup><sub>rp</sub><br>%<sub>0</sub></th><th>Reference</th></tr></thead><tbody><tr><td rowspan="5">1,2,4-TMB</td><td>Nitrate</td><td>Enrichment culture</td><td>-0.4 ±<br>0.2</td><td>1.004 ±<br>0.002</td><td>-12 ± 3</td><td>1.168 ±<br>0.049</td><td>29 ± 17</td><td>12 ±<br>11</td><td>This study</td></tr><tr><td>Iron</td><td>Enrichment culture</td><td>-0.3 ±<br>0.1</td><td>1.002 ±<br>0.001</td><td>-14 ± 1</td><td>1.202 ±<br>0.017</td><td>52 ± 21</td><td>24 ±<br>26</td><td>This study</td></tr><tr><td>Manganese</td><td>Enrichment culture</td><td>-0.2 ±<br>0.1</td><td>1.001 ±<br>0.001</td><td>-19 ± 2</td><td>1.295 ±<br>0.040</td><td>120 ±<br>70</td><td>50 ±<br>45</td><td>This study</td></tr><tr><td>Sulfate</td><td>Enrichment culture</td><td>-0.6 ±<br>0.1</td><td>1.006 ±<br>0.001</td><td>-18 ± 2</td><td>1.276 ±<br>0.039</td><td>29 ± 6</td><td>13 ± 9</td><td>This study</td></tr><tr><td>Nitrate</td><td>Enrichment culture</td><td>-1.4 ±<br>0.2</td><td>1.013 ±<br>0.002</td><td>-31 ± 5</td><td>1.592 ±<br>0.152</td><td>23 ± 4</td><td>10 ± 7</td><td>This study</td></tr><tr><td rowspan="5">1,3,5-TMB</td><td>Iron</td><td>Enrichment culture</td><td>-0.3 ±<br>0.1</td><td>1.003 ±<br>0.001</td><td>-20 ± 3</td><td>1.316 ±<br>0.062</td><td>58 ± 23</td><td>27 ±<br>23</td><td>This study</td></tr><tr><td>Manganese</td><td>Enrichment culture</td><td>-0.4 ±<br>0.1</td><td>1.004 ±<br>0.001</td><td>-22 ± 3</td><td>1.359 ±<br>0.066</td><td>50 ± 16</td><td>22 ±<br>18</td><td>This study</td></tr><tr><td>Sulfate</td><td>Enrichment culture</td><td>-0.8 ±<br>0.3</td><td>1.008 ±<br>0.003</td><td>-11 ± 3</td><td>1.152 ±<br>0.048</td><td>14 ± 6</td><td>6 ± 5</td><td>This study</td></tr><tr><td>Nitrate</td><td>Enrichment culture</td><td>-0.1 ±<br>0.1</td><td>1.001 ±<br>0.001</td><td>-12 ± 4</td><td>1.168 ±<br>0.066</td><td>107 ±<br>91</td><td>53 ±<br>68</td><td>This study</td></tr><tr><td>Iron</td><td>Enrichment culture</td><td>-0.6 ±<br>0.1</td><td>1.006 ±<br>0.001</td><td>-10 ± 1</td><td>1.136 ±<br>0.015</td><td>16 ± 2</td><td>7 ± 7</td><td>This study</td></tr><tr><td rowspan="5">1,2,3-TMB</td><td>Manganese</td><td>Enrichment culture</td><td>-0.4 ±<br>0.1</td><td>1.004 ±<br>0.001</td><td>-7 ± 3</td><td>1.092 ±<br>0.043</td><td>16 ± 9</td><td>7 ± 5</td><td>This study</td></tr><tr><td>Sulfate</td><td>Enrichment culture</td><td>-1.1 ±<br>0.1</td><td>1.010 ±<br>0.001</td><td>-15 ± 10</td><td>1.220 ±<br>0.178</td><td>14 ± 9</td><td>6 ± 5</td><td>This study</td></tr><tr><td rowspan="10">Toluene</td><td>Nitrate</td><td>Azoarcus sp. EbN1 (n = 3)</td><td>-3.2 ±<br>0.5</td><td>1.023 ±<br>0.004</td><td>-51 ± 7</td><td>1.689 ±<br>0.163</td><td>16 ± 3</td><td>6 ± 2</td><td>Vogt et al. (2008)</td></tr><tr><td>Nitrate</td><td>Azoarcus sp. T (n = 3)</td><td>-5.8 ±<br>0.3</td><td>1.042 ±<br>0.002</td><td>-79 ± 1</td><td>2.725 ±<br>0.067</td><td>14 ± 1</td><td>5 ± 2</td><td>Vogt et al. (2008)</td></tr><tr><td>Nitrate</td><td>DD-Anox1</td><td>-3.2 ±<br>0.4</td><td>1.023 ±<br>0.003</td><td>-73 ± 9</td><td>2.404 ±<br>0.416</td><td>23 ± 4</td><td>9 ± 4</td><td>Dorer et al. (2016)</td></tr><tr><td>Nitrate</td><td>Georgfuchsia toluolica strain G5G6</td><td>-2.6 ±<br>0.1</td><td>1.019 ±<br>0.001</td><td>-58 ± 4</td><td>1.866 ±<br>0.111</td><td>22 ± 2</td><td>9 ± 3</td><td>Dorer et al. (2016)</td></tr><tr><td>Nitrate</td><td>Thauera aromatica</td><td>-2.7 ±<br>0.1</td><td>1.019 ±<br>0.001</td><td>-35 ± 14</td><td>1.389 ±<br>0.216</td><td>13 ± 5</td><td>5 ± 2</td><td>Vogt et al. (2008)</td></tr><tr><td>Nitrate</td><td>Geobacter metallireducens, solid Fe(III)</td><td>-1.3 ±<br>0.1</td><td>1.009 ±<br>0.001</td><td>-37.2 ±<br>1</td><td>1.424 ±<br>0.016</td><td>29 ± 2</td><td>11 ± 4</td><td>Tobler et al. (2008)</td></tr><tr><td>Nitrate</td><td>Geobacter metallireducens, solid Fe(III)-citrate</td><td>-3.6 ±<br>0.7</td><td>1.026 ±<br>0.005</td><td>-98.4 ±<br>3</td><td>4.699 ±<br>0.530</td><td>27 ± 5</td><td>10 ± 5</td><td>Tobler et al. (2008)</td></tr><tr><td>Nitrate</td><td>Georgfuchsia toluolica strain G5G6</td><td>-1.7 ±<br>0.2</td><td>1.012 ±<br>0.001</td><td>-76 ± 4</td><td>2.551 ±<br>0.208</td><td>45 ± 6</td><td>10 ± 4</td><td>Dorer et al. (2016)</td></tr><tr><td>Nitrate</td><td>Aromatoleum aromaticum strain EbN1</td><td>-1.2 ±<br>0.2</td><td>1.008 ±<br>0.001</td><td>-18 ± 3</td><td>1.168 ±<br>0.033</td><td>15 ± 4</td><td>17 ± 6</td><td>Dorer et al. (2016)</td></tr><tr><td>Nitrate</td><td>DD-Anox1</td><td>-1.8 ±<br>0.3</td><td>1.013 ±<br>0.002</td><td>-36 ± 7</td><td>1.404 ±<br>0.110</td><td>20 ± 5</td><td>6 ± 2</td><td>Dorer et al. (2016)</td></tr><tr><td>Nitrate</td><td>Georgfuchsia toluolica strain G5G6</td><td>-0.7 ±<br>0.2</td><td>1.005 ±<br>0.001</td><td>-17 ± 3</td><td>1.157 ±<br>0.032</td><td>24 ± 8</td><td>8 ± 3</td><td>Dorer et al. (2016)</td></tr><tr><td>Nitrate</td><td>Desulfosarcina cetonica (n = 3)</td><td>-2.4 ±<br>0.2</td><td>1.017 ±<br>0.001</td><td>-74 ± 6</td><td>2.445 ±<br>0.287</td><td>31 ± 4</td><td>9 ± 4</td><td>Vogt et al. (2008)</td></tr><tr><td>Nitrate</td><td>Strain TRM1</td><td>-2.0 ±<br>0.2</td><td>1.014 ±<br>0.001</td><td>-66 ± 19</td><td>2.119 ±<br>0.682</td><td>33 ± 10</td><td>12 ± 5</td><td>Vogt et al. (2008)</td></tr><tr><td rowspan="5">Ethylbenzene</td><td>Iron</td><td>Georgfuchsia toluolica strain G5G6</td><td>-2.4 ±<br>0.2</td><td>1.010 ±<br>0.001</td><td>-65 ± 13</td><td>2.857 ±<br>1.061</td><td>27 ± 6</td><td>13 ± 5</td><td>Dorer et al. (2016)</td></tr><tr><td>Manganese</td><td>Aromatoleum aromaticum strain EbN1</td><td>-1.6 ±<br>0.1</td><td>1.006 ±<br>0.000</td><td>-50 ± 4</td><td>2.000 ±<br>0.160</td><td>31 ± 3</td><td>2 ± 1</td><td>Dorer et al. (2016)</td></tr><tr><td>Manganese</td><td>DD-Anox1</td><td>-1.3 ±<br>0.4</td><td>1.005 ±<br>0.002</td><td>-78 ± 10</td><td>4.545 ±<br>2.066</td><td>60 ± 20</td><td>23 ± 7</td><td>Dorer et al. (2016)</td></tr><tr><td>Manganese</td><td>Georgfuchsia toluolica strain G5G6</td><td>-1.7 ±<br>0.2</td><td>1.007 ±<br>0.001</td><td>-53 ± 10</td><td>2.128 ±<br>0.453</td><td>31 ± 7</td><td>26 ± 8</td><td>Dorer et al. (2016)</td></tr><tr><td>Manganese</td><td>Strain TRM1</td><td>-0.7 ±<br>0.1</td><td>1.006 ±<br>0.001</td><td>-25 ± 3</td><td>1.333 ±<br>0.053</td><td>36 ± 7</td><td>50 ±<br>22</td><td>Herrmann et al.<br>(2009)</td></tr><tr><td rowspan="3">o-Xylene</td><td>Sulfate</td><td>Consortium</td><td>-2.3 ±<br>0.4</td><td>1.019 ±<br>0.003</td><td>-41 ± 9</td><td>1.695 ±<br>0.259</td><td>18 ± 5</td><td>26 ± 8</td><td>Herrmann et al.<br>(2009)</td></tr><tr><td>Sulfate</td><td>Desulfosarcina ovata</td><td>-2.3 ±<br>0.1</td><td>1.019 ±<br>0.001</td><td>-50 ± 6</td><td>2.000 ±<br>0.240</td><td>22 ± 3</td><td>15 ± 9</td><td>Herrmann et al.<br>(2009)</td></tr><tr><td>Sulfate</td><td>Consortium</td><td>-2.7 ±<br>0.2</td><td>1.022 ±<br>0.002</td><td>-48 ± 5</td><td>1.923 ±<br>0.185</td><td>18 ± 2</td><td>7 ± 4</td><td>Herrmann et al.<br>(2009)</td></tr><tr><td rowspan="2">p-Xylene</td><td>Sulfate</td><td>Consortium</td><td>-1.2 ±<br>0.3</td><td>1.010 ±<br>0.002</td><td>-19 ± 3</td><td>1.235 ±<br>0.046</td><td>16 ± 5</td><td>9 ± 5</td><td>Herrmann et al.<br>(2009)</td></tr><tr><td>Sulfate</td><td>Consortium</td><td>-1.2 ±<br>0.3</td><td>1.010 ±<br>0.002</td><td>-19 ± 3</td><td>1.235 ±<br>0.046</td><td>16 ± 5</td><td>9 ± 5</td><td>Herrmann et al.<br>(2009)</td></tr></tbody></table>
exhibited dense microbial colonization. Both studies concluded that when microorganisms are attached to solid phases or form biofilms, mass transfer limitations of the substrate into the cells became rate-limiting and therefore mask isotope fractionation. Since we used limonite and cryptomelane respectively as iron and manganese sources, we could postulate that the same phenomenon occurred in our experiments. Extending this assumption, the low solubility of minerals may lead to mass transfer limitations, as they predominantly exist in the solid phase in anoxic groundwater, therefore allowing microorganisms to attach and form biofilms. The BSS commitment to catalysis during the fumarate addition could also be responsible for isotope fractionation masking via, therefore reducing AKIE values (Northrop, 1981; Elsner, 2010). During degradation under denitrifying conditions, nitrite toxicity may have reduced BSS activity, lowering the commitment to catalysis (Haner et al., 1997). This partial inhibition could explain the higher fractionation values, as decreased enzyme activity reduces isotope fractionation masking, leading to the highest enrichment observed among all conditions.

In order to circumvent masking of isotope fractionation, $\Lambda_C^H$ per compound and per condition were computed using enrichment factors for carbon and hydrogen. Indeed, since non-fractionating steps mask in the same extent both carbon and hydrogen isotopes, $\Lambda_C^H$ are supposed to be unaffected by masking (Elsner et al., 2005). This $\Lambda_C^H$ calculation method was preferred over the conventional dual-isotope plot due to the significant sample loss encountered during the study. Despite following the same fumarate addition pathway, $\Lambda_C^H$ can differ depending on the condition. Slopes were significantly different between each isomer, with 1,2,4-TMB having overall the highest value, ranging from 29 ± 17 to 120 ± 70, while 1,2,3-TMB presented overall the lowest ones except in nitrate-reducing condition. Very high values were recorded for 1,2,4-TMB in manganese-reducing condition (120 ± 70) and 1,2,3-TMB in denitrifying conditions (107 ± 91), and were considered with precaution. The $\Lambda_C^H$ seems to follow the same trend as the biodegradability of the trimethylbenzene isomer, with the most biodegradable isomer exhibiting highest $\Lambda_C^H$. Such differences in $\Lambda_C^H$ could also reflect the occurrence of different reaction mechanisms, but this hypothesis has been discarded in regard to the degradation products formed and to observations reported in previous studies (Beller, 2000; Morasch et al., 2004). Since no $\Lambda_C^H$ are available for TMBs, $\Lambda_C^H$ was preferred to compare our values to the literature for other alkylbenzenes degradation via fumarate addition. With the exception of 1,2,4-TMB under manganese-reducing conditions and 1,2,3-TMB under denitrifying conditions, $\Lambda_{Tp}^{H}$ aligned with values observed for other alkylbenzenes, further confirming the occurrence of fumarate addition. Difference of $\Lambda_{Tp}^{H}$ for the same reaction can differ depending on the nature of rate-determining step (Elsner, 2010). Kümmel et al. (2013) concluded that subtle $\Lambda_C^H$ differences could be attributed to slight differences in the phylogeny of the BSS, leading to different reaction paths. Finally, as observed by Pinel-Cabello et al. (2024) for toluene anaerobic biodegradation, high $\Lambda_C^H$ variability can be exhibited among bacteria following the same catabolic pathway. A more detailed study of the non-fractionating steps inducing variability in isotope fractionation of TMBs could eventually be performed using labelled substrate, with heavy isotope occupying every reactive position, but it fell outside the scope of this study.

## 4. Conclusion

This study revealed the potential biodegradability of TMB isomers under anaerobic conditions in groundwater. Collected lines of evidence tend to confirm that each TMB isomer undergoes biodegradation under

denitrifying, iron-, manganese- and sulfate-reducing conditions. Concentration measurement and metabolite identification helped to confirm that fumarate addition is the common pathway to TMBs and more generally to alkylbenzenes. Despite a low enrichment in carbon and moderate in hydrogen isotopes in comparison with other studies, results obtained via multi-elemental compound specific isotope analysis fell in accordance with the degradation via fumarate addition. Moreover, low carbon and hydrogen fractionation, associated with their strong variability, tends to suggest that non-fractionating steps provoke isotope fractionation masking. This study is the first to produce enrichment factors in carbon and hydrogen for TMBs, which can ultimately help at field site to quickly assess if biodegradation occurs or not. Finally, our results suggest that TMBs can be used as conservative tracers for biodegradation of alkylbenzenes, but with precautionary care. Indeed, TMB degradation will likely occur only when it will be biologically favorable for indigenous strains, in other words only when more biodegradable carbon source (toluene, xylene, etc.) have been depleted, and therefore serve in the meantime to trace anaerobic removal processes.

## CRediT authorship contribution statement

Ouassim Boukaroum: Writing – original draft, Methodology, Investigation, Formal analysis, Data curation, Conceptualization. Carine Demelas: Methodology, Data curation, Conceptualization. Laurent Vassalo: Methodology, Investigation, Data curation. Patrick Höhener: Funding acquisition, Conceptualization, Supervision, Methodology.

## Declaration of competing interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

## Acknowledgments

This work is funded by the French National Research Agency (ANR) through grant ANR-18-CE04-0004, project DEGiSIVE and by the Doctoral College Environmental Sciences (ED251).

## Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.jconhyd.2025.104666.

## Data availability

Data will be made available on request.

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