# Electrochemical oxidation of polycyclic aromatic hydrocarbons in soil co-contaminated with arsenic

Kim Johansson, Azariel Ruiz-Valencia, Ivan Carabante, Timothy M Vogel, Erkki Lindberg, Jurate Kumpiene

► To cite this version:

Kim Johansson, Azariel Ruiz-Valencia, Ivan Carabante, Timothy M Vogel, Erkki Lindberg, et al. Electrochemical oxidation of polycyclic aromatic hydrocarbons in soil co-contaminated with arsenic. Journal of Environmental Management, 2025, 388, pp.125978. 10.1016/j.jenvman.2025.125978. hal-05447936

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Research article

# Electrochemical oxidation of polycyclic aromatic hydrocarbons in soil co-contaminated with arsenic

Kim Johansson<sup>a, *</sup> <sup>□</sup>, Azariel Ruiz-Valencia<sup>b, □</sup>, Ivan Carabante<sup>a</sup>, Timothy M. Vogel<sup>b</sup>, Erkki Lindberg<sup>c</sup>, Jurate Kumpiene<sup>a</sup>

<figure><img src="image_3.png" /></figure>

## ARTICLE INFO

**Keywords:**

PAH

soil solution

In situ soil remediation

Microbiological analyses

## ABSTRACT

The wood preservation industry has contaminated numerous sites with polycyclic aromatic hydrocarbons (PAH). The aim of this study was to investigate the use of electrochemical oxidation (EO), without chemical additives, for *in situ* degradation of PAHs in soil co-contaminated with arsenic. Two 1 m³ boxes, each containing a 20 cm layer of contaminated sand overlaid with 20 cm of peat, were equipped with iron electrodes and placed outdoors. EO was applied using a pulsating direct current with alternating polarity, and its impact on PAH concentrations in soil and soil solution, as well as on the associated microbial community was assessed. Soil solution was sampled from the boxes over two seasons and analysed for PAH16 concentrations, showing an average decrease of 82 % by the end of the second season. This reduction was mainly observed in the medium and high molecular weight PAH fractions, suggesting that EO can effectively degrade more recalcitrant PAH compounds. The least reduction was seen in the low molecular weight PAH16, likely due to the replenishment from PAHs sorbed to the soil. Soil samples were taken from 12 different locations within the boxes, showing that PAH16 concentrations significantly decreased in 10 out of 12 sampling points, with a greater average reduction in sand (84 % decrease) compared to peat (69 % decrease). Microbial analyses of the soil samples revealed no significant changes in DNA concentrations across all taxa over time, with *Gammaproteobacteria* remaining the most abundant microorganisms in all samples, suggesting its ability to persist complex contamination.

## 1. Introduction

In the past, creosote was commonly used as a wood preservative, containing high concentrations of polycyclic aromatic hydrocarbons (PAH). Furthermore, along with creosote, CCA (copper, chromium, arsenic) was also used in wood treatment. This industrial use has led to soil contamination at numerous sites that still require remediation today, as PAH are persistent, toxic and cancerogenic, posing significant risks to both human health and the environment (Jones and Peace, 1989; Lee et al., 2017; White and Claxton, 2004). PAH are nonpolar compounds consisting of two or more aromatic rings (Connell, 1997) and many of the properties are linked to their molecular weight. Increasing molecular weight leads to higher lipophilicity and lower water solubility causing PAH to strongly sorb to soil organic matter (OM) (Chen et al., 2017; Koelmans et al., 2006; Okere, 2012; Wilson and Jones, 1993). Sum of 16 PAH are listed as high priority pollutants

worldwide (ATSDR, 2011; Mumtaz et al., 1996). In Sweden, they are divided into three subcategories, PAH-L, PAH-M, and PAH-H, based on their molecular weight (low, medium, high, respectively) with distinct guideline values for contaminated soil for each group (SEPA, 2009).

Biodegradation of PAHs, which occurs through enzymatic processes and microbial oxidation (Gan et al., 2009), is often preferred when selecting *in situ* soil remediation techniques. However, due to the persistence of PAHs, this process is often slow, especially in colder climates (Eriksson et al., 2003). Additionally, older contamination at historic sites may be dominated by more recalcitrant PAH fractions (i.e., higher molecular weight compounds), limiting their bioavailability for microbial degradation. Although the use of creosote and CCA in Sweden ceased decades ago, high concentrations of these chemicals can still be found co-present in soil at former wood impregnation sites. Toxic levels of As, Cr, and Cu in soil and groundwater can further inhibit bacterial growth and PAH degradation. Therefore, techniques that accelerate

<footnote><sup>a</sup> Waste Science and Technology, Luleå Technical University, Luleå, 971 87, Sweden</footnote>

<footnote><sup>b</sup> Ecologie Microbienne, UMR CNRS 5557, UMR INRAE 1418, Universite Claude Bernard Lyon 1, Villeurbanne, France</footnote>

<footnote><sup>c</sup> EkoGrid Oy, Nuijamiestentie 7, Helsinki, 00400, Finland</footnote>

<footnote>* Corresponding author.</footnote>

<footnote>E-mail address: kim.johansson@ltu.se (K. Johansson).</footnote>
PAH degradation despite soil toxicity should be considered, either as alternatives to or in complement with bioremediation.

Electrokinetic remediation (EK) is a technology available for PAH degradation in situ (Virkutyte et al., 2002). EK has traditionally been used to remove metal contaminants from soil (Virkutyte et al., 2002; Wen et al., 2021). Later, the method was extended to the removal of a broad range of organic compounds (Lim et al., 2016; Sabino da Silva et al., 2017), and more recently, it has been applied for the degradation of organic compounds in soil through electrochemical oxidation (Gan et al., 2009; Huang et al., 2012). The latter processes of degradation of organic compounds within soil, rather than merely desorbing and transporting them through the soil for subsequent removal, has considerably increased the potential of using electrical field for reme-diation of PAH contaminated soil in situ (Proietto et al., 2023, 2024). It has been shown that even at very low electric field strength (0.15 V cm⁻¹), organic contaminants can be simultaneously desorbed, mobilised and degraded in situ (Proietto et al., 2023). It has been also suggested that EO occurs in symbiosis with microbiological degradation, which further enhances the overall remediation effect (Fan et al., 2007; Guo et al., 2014; Kim et al., 2005). However, the use of surface-active sub-stances and oxidants have been suggested as necessary for enhancing the desorption from soil OM or for the pre-oxidation of organic contami-nants, thereby increasing the effectiveness of EO remediation (Alshawabkeh et al., 2004; Saichek and Reddy, 2005; Yeom et al., 1996). Recent studies (2017–2020) on EO remediation of both organic and inorganic contaminants used additives (Abou-Shady et al., 2023). For example, hydrogen peroxide (H₂O₂) used for the oxidation of PAH in a process known as chemical oxidation, is a common additive in remedi-ation of PAH contaminated soil (Gan et al., 2009; Valderrama et al., 2009). However, this implies environmental risks related with poten-tially negative impacts on soil biota (Valderrama et al., 2009) and safe handling of chemicals, especially in full scale applications (Yap et al., 2011).

The challenge of increasing PAH availability for degradation could be addressed by modifying the method of supplying electrical current to the soil. By applying low-voltage current in a pulsating mode and reversed polarity, rather than a continuous electrical flow, the effective energy delivered through the soil may increase, enhancing the desorp-tion of PAHs from soil particles without the need for additional chem-icals. Furthermore, if electrodes are made of iron (Fe), their corrosion can introduce Fe(II)/Fe(III) into the soil. This process, combined with the potential formation of reactive oxygen species through water hy-drolysis, could likely enhance the production of hydroxyl radicals (·OH) via a Fenton-like reaction, which are important for the chemical oxidation (Liu et al., 2023; Nong et al., 2015; Sun et al., 2021).

In this study, electrochemical oxidation was applied for degradation of PAH16 to contaminated soil and soil solution. The aim of this study was to use electrochemical oxidation induced by low voltage pulsing current with alternating polarity in an outdoor experiment without additional chemicals and to evaluate its impact on PAH16 degradation, by i) monitoring changes in contaminant concentration in soil and soil solution and ii) analysing changes in microbial community structure as affected by the treatment. This technique is intended to be applied in situ without soil excavation for consequent transportation to external facil-ities. This is the second part of a study on creosote (PAH)-arsenic (CCA) co-contaminated soil, while the first part, which focused on the impact of electrochemical oxidation on arsenic mobility, is described in Johansson et al. (2024).

# 2. Materials and methods

## 2.1. Soil samples

In May 2021, polluted soil was sampled from a former industrial site in Limmared, Sweden, where creosote and CCA were used to treat wood railway sleepers. At the sampling site, the top layer of soil, rich in

organic matter, was designated as "peat." Beneath the peat lay a layer of sandy soil, referred to as "sand." Approximately 400 kg of each soil type were excavated and transported to the university for experimental purposes. Before the experiments, rocks and coarse particles larger than 10 cm were removed, and the peat and sand were separately homoge-nized using a cement mixer.

## 2.2. Experimental setup

Two 1 m³ (1m × 1m x 1m) boxes of polyethylene (box 1 and box 2, used as replicates) were filled with homogenized soil similarly to its structure in field, a bottom layer with 20 cm of sand, and a 20 cm layer of peat on top. Two zerovalent iron (ZVI) electrodes (1 m, Ø 1 cm) were placed vertically in the opposite corners of each box and connected to a power control unit (Ekogrid, Finland). The current direction was reversed once a week to enable each electrode to alternate between an anodic and cathodic state. The voltage potential was 0.025 V cm⁻¹ and the current was pulsed at frequencies ranging from 500 to 1000 pulses per second throughout the duration of the experiment. The boxes were covered with a transparent plastic film to prevent precipitation from entering the boxes, while still allowing air to enter. Fresh tap water was added periodically to the boxes to maintain the soil humidity and to simulate soil solution. In the centre of each box, a polyethylene perforated groundwater sampling pipe was installed to facilitate the sampling of soil solution. The water level was monitored through the pipe, with the level maintained at 10 cm below the soil surface. This ensured that the sand layer was fully saturated, along with the bottom half of the peat layer. Additionally, 24 porewater samplers (9 cm, Ø 4.5 mm, pore size 0.12–0.18 mm, MacroRhizon, Rhizosphere Research Products, the Netherlands) were installed, six in each soil layer (12 per box), to monitor changes in soil solution pH, redox potential, electrical con-ductivity and elemental concentrations. Two smaller containers (ca 20 kg each) were filled with soil in the same manner as the boxes, equipped with porewater samplers, stored under the same conditions and used as controls (untreated soil). The experimental setup is illustrated in Fig. S1.

The experiment was conducted outdoors over two summer seasons under natural light and temperature conditions in Luleå, Sweden (65°37'11.1"N 22°08'37.4"E). The first season lasted 4.5 months, from June to October 2021, and the second season ran the following year for 2.5 months, from July to mid-September 2022.

## 2.3. Sampling and chemical analyses

Initial soil samples of both sand and peat were collected after ho-mogenization and before placing the soil in the experiment boxes for PAH16 analyses. At the end of the first season, a total of twelve soil samples (50 g each) were collected, three from the sand layer and three from the peat layer in each box. The procedure was repeated at the end of the second season. These samples were taken from areas around the electrodes (E1, E2) and in the centre (C) of the boxes (Fig. S1). In par-allel, untreated soil samples from each soil layer were collected at the end of each season. All samples were immediately placed in diffusion-tight bags and stored in cool and dark conditions until analysed for PAH16 with Gas Chromatography-Mass Spectrometry (GC-MS) at an accredited laboratory (ALS Scandinavia) following US EPAs methods 8270D and 8000D.

Porewater samples for the analysis of physicochemical properties and elemental concentrations were collected twice a week during the first season (a total of 22 sampling occasions) and every second or third week during the second season (a total of four sampling occasions). Samples were immediately analysed for pH (WHT pH 340/SET-1), electric conductivity (EC) (Radiometer Copenhagen CDM 210 conduc-tivity meter) and redox potential (Eh) (Radiometer Copenhagen PHM 95 pH/ion meter). The remaining samples were acidified with concentrated nitric acid and stored at 4 °C prior to arsenic (As) and iron (Fe) analysis by inductively coupled plasma – optical emission spectrometry (ICP-
OES, PerkinElmer optima 8300). Soil solution samples of 250 mL were taken once a month from the central pipes and analysed for their PAH16 concentrations with GC-MS after decantation at accredited laboratory (ALS Scandinavia).

## 2.4. Microbiological analyses

For microbiological tests, a small portion (0.5 g) of the collected peat and sand was used for DNA extraction. These samples were taken at the beginning of the first season (A1), at the end of the first season (A2), at the beginning of the second season (B1), and at the end of the second season (B2). The bacterial DNA was extracted from sand or peat with the kit for DNA NucleoSpin Soil (Macherey-Nagel, Germany), according to manufacturer instructions. The V4 region from the 16S rRNA gene was amplified using the Illumina primers ILL_515Fmod and ILL_806Rmod. Unique adapters were added to each sequence to identify the different samples. After each PCR step, the PCR products were cleaned using magnetic AMPure XP Beads (Beckman, USA). Finally, the amplified sequences were pooled together in equimolar concentrations and loaded into a MiSeq sequencer following the manufacture protocol (Illumina, UK). The obtained sequences were denoised using DADA2 to obtain amplicon sequence variants. The taxonomy for each Amplicon Sequence Variant (ASV) was annotated against the SILVA NR99 v138.1 database.

## 2.5. Statistics

Differential expression analysis for the ASVs and the counts was performed with the DESeq2 package from R (version R 4.4.1) in Rstudio (Love et al., 2014). Principal component analysis was performed using Bray-Curtis distances to visually display variability samples. Axes variances were calculated without zeroing negative eigenvalues. Correlation plots were done with the package corrplot (0.95).

# 3. Results and discussion

## 3.1. Soil properties

At the sampling site in Limmared, free-phase creosote oil was found in the soil. Since PAH compounds are the primary constituents of creosote oil used in the wood treatment industry (Mueller et al., 1989), high concentrations of PAHs are commonly found at creosote-contaminated sites (Simpanen et al., 2016; Zapf-Gilje and Patrick, 2001). The PAH16 concentration in a homogenized composite sample collected prior to the experiment was two orders of magnitude higher in peat than in sand, most likely due to the significantly higher organic matter (OM) content in peat (Table 1). Soil organic matter can strongly sorb PAH, reducing its mobility and bioavailability, which makes PAHs less prone to leaching but more persistent in the environment (Ukalska-Jaruga and Smreczak, 2020; Zeng et al., 2019). The concentrations of low, medium, and high molecular weight PAH fractions in peat substantially exceeded the Swedish generic guideline values for contaminated soil intended for less sensitive land use (industrial sites) (SEPA, 2009). The medium

molecular weight PAH constituted the dominant fraction of PAH in peat (66 % of PAH16). It is noteworthy that PAH concentration in the initial sandy soil sample was lower than expected, which could be attributed to the difficulties encountered in homogenizing the soil with free-phase contamination.

The pH and Eh were similar for both sand and peat, while the electrical conductivity (EC) was significantly higher in peat than in sand (Table 1).

## 3.2. Changes in soil solution pH, redox potential and electrical conductivity

Measurements of pH in porewater samples showed no significant variation based on the location of the samplers within the treatments. This suggests that the power unit settings, specifically the application of an alternating current direction, effectively prevented the formation of a pH gradient, despite the absence of chemical additives. Previously, the use of chemical additives has been proposed as a necessity to keep pH stable and to desorb PAH from soil in EO setups (Abou-Shady et al., 2023; Cameselle and Gouveia, 2018; Reddy and Saichek, 2003). However, there were some oscillations in pH, redox potential and electrical conductivity over time. Porewater pH values slightly dropped during the second sampling season by one pH-unit from the initial pH ca 6.0-7.5 to the final pH ca 5.0-6.5 (Figure S2 C,D, Johansson et al., 2024). Additionally, by the end of the first season, around sampling days 99-107, a decrease in redox potential was observed, dropping from 125 ± 9 mV to 3 ± 32 mV (Figure S2 A,B, Johansson et al., 2024). During the same time, the daily average outdoor temperature dropped by 9 °C (from 14.8 to 5.8 °C), and with that, the evaporation of soil water decreased. prolonged season of water saturated conditions can lead to oxygen depletion, especially in organic matter rich soil, which was the most likely reason for the observed drop in redox values.

The electrical conductivity of porewater in sand was, on average, higher than that in peat over the course of the seasons (339 ± 151 µS cm⁻¹ and 203 ± 72 µS cm⁻¹, respectively). However, by the end of the experiment, conductivity levels in both soil types equalized, converging to a similar range of 117-146 µS cm⁻¹ (Figure S2 E,F).

Variations in pH have shown to not affect PAH degradation in soil and sediments (Kästner et al., 1998; Shiaris, 1989). Conductivity, on the other hand could affect PAH degradation, since PAH solubility can decrease with increasing salinity (Shiaris, 1989).

## 3.3. Impact of electrooxidation on PAH in soil solution

A comparison of PAH16 concentrations in the soil solution from the beginning of the first season to the end of the second season shows an overall reduction of 82 % (Fig. 1). At the beginning of the experiment, the largest group of PAHs in soil solution was medium molecular weight PAHs (PAH-M), comprising 74 % of the total dissolved PAHs. This was followed by high molecular weight PAHs (PAH-H), which accounted for 26 % of the total PAH, while the low molecular weight PAH fraction (PAH-L) was negligible, making up only 0.2 % of the total PAH. This

<caption>**Table 1**</caption>

Main properties of the two studied soils at the beginning of the experiment along with the Swedish EPA guideline value (GV) for PAH in contaminated soil with less sensitive land use (industrial sites) (SEPA, 2009).

<table><thead><tr><th>Soil</th><th>pH</th><th>Eh, mV</th><th>EC, µS cm<sup>-1</sup></th><th>Organic matter<sup>a</sup>, %</th><th>PAH16, mg kg<sup>-1</sup></th><th>PAH-L<sup>b</sup>, mg kg<sup>-1</sup></th><th>PAH-M<sup>c</sup>, mg kg<sup>-1</sup></th><th>PAH-H<sup>d</sup>, mg kg<sup>-1</sup></th></tr></thead><tbody><tr><td>Sand</td><td>5.35 ± 0.07</td><td>380 ± 11</td><td>82 ± 10</td><td>2.2 ± 0.2</td><td>14.4 ± 4.3</td><td>0.69 ± 0.21</td><td>5.51 ± 1.65</td><td>8.17 ± 2.45</td></tr><tr><td>Peat</td><td>5.09 ± 0.03</td><td>334 ± 9</td><td>409 ± 119</td><td>88.9 ± 2.1</td><td>5160 ± 1550</td><td>841 ± 252</td><td>3380 ± 1014</td><td>940 ± 282</td></tr><tr><td>SEPA GV</td><td></td><td></td><td></td><td></td><td></td><td>15</td><td>20</td><td>10</td></tr></tbody></table>

<footnote>a Measured as loss on ignition at 550 °C.</footnote>

<footnote>b PAH-L: naphthalene, acenaphthene, and acenaphthylene.</footnote>

<footnote>c PAH-M: fluorene, phenanthrene, anthracene, fluoranthene, and pyrene.</footnote>

<footnote>d PAH-H: Benzo(a)anthracene, chrysene, benzo(b)fluoranthene, benzo(k)fluoranthene, benzo(a)pyrene, indeno(1,2,3-c,d)pyrene, dibenzo(a,h)anthracene, and benzo(g,h,i)perylene.</footnote>
<figure><img src="image_4.png" /><figcaption>**Fig. 1.** Changes in PAH concentrations (µg L⁻¹) in soil solution during the two seasons. Average of two replicates (boxes). S1-first season, S2-second season.</figcaption></figure>

distribution is not unexpected, as PAH-L in historical sites have had time to leach, degrade, or volatilize over time. PAH-L compounds are the most soluble and easily degradable, resulting in relatively low concentrations in aged soils compared to the more recalcitrant PAH fractions with heavier molecules.

Sum PAH16 concentration in soil solution peaked during the second sampling in July of 2021 (Fig. 1). The relative fraction that increased the most during the peak concentrations was PAH-L (from 0.2 to 8.4 % of total PAH). This peak may be attributed to the desorption of weakly adsorbed PAHs from soil particles as a result of the EO treatment with its pulsating settings, leading to their rapid release into the soil solution. The desorption process is desired during the soil remediation as increased PAH desorption increases their availability for degradation, where electrochemical oxidation is believed to take place. Wen et al. (2023) showed that in an EO treatment, the transportation of free radicals and their mixing with PAH were enhanced by a pulsed DC current, leading to increased PAH degradation. Generation of reactive oxygen species (ROS), particularly hydroxyl radicals (·OH) would then allow for the PAH to be oxidized in a stepwise manner (Andreottola and Ferrarese, 2008).

The fractions of PAH that decreased the most in soil solution over two seasons were PAH-M (90 % decrease) and PAH-H (63 % decrease) (Fig. 2). Due to these changes in the fractional distribution of PAH compounds, the dominant PAH fraction by the end of the experiment shifted to PAH-H (54 %), followed by PAH-M (44 %) and PAH-L (2 %).

<figure><img src="image_5.png" /><figcaption>**Fig. 2.** Changes in relative distribution of PAH groups (PAH-L, PAH-M, PAH-H) during the experiment. S1-first season, S2-second season.</figcaption></figure>

At the beginning of the experiment, fluoranthene, pyrene, benz(a) anthracene, chrysene, and benzo(b)fluoranthene accounted for 92 % of the PAH16 in the soil. These are all common ingredients in creosote (Mueller et al., 1989; Northcott and Jones, 2001). At the end of the first season, the concentration of those five most abundant PAH16 compounds in the soil solution decreased by 44–93 %, indicating that EO can efficiently target contaminants from creosote.

It was expected that the PAH-L concentrations would decrease the most as these compounds are more water-soluble and have lower affinity for organic matter compared to PAH with medium and high molecular weight (Eldos et al., 2022) and therefore are easiest to degrade both chemically (Jonsson et al., 2007) and biologically (Silva et al., 2009). Also, as PAH-L compounds are more volatile than those with higher molecular weight (Ma et al., 2006), some losses of PAH-L from soil might occur through volatilization (Wang et al., 2016). Since the treatment boxes were not airtight, emissions from the soil to air could not be quantified, meaning that the decline in PAH could be due to the EO treatment, biodegradation or volatilization. Nevertheless, the absolute values of the PAH-L fraction were very low (0.6–19 µg/L), except for the peak during the second sampling of the study, and did not change significantly by the end of the experiment (2.9 ± 0.7 µg/L) compared to its starting values (2.3 ± 0.9 µg/L). Acenaphthene was the dominant compound in the PAH-L group, comprising an average of 80 % of the PAH-L fraction. Transformation, especially biological degradation, of 3-ring PAHs such as acenaphthene is generally slower than that of 2-ring naphthalene (Mueller et al., 1989), which might lead to relative enrichment of acenaphthene compared to naphthalene. Thus, if the biodegradation of PAH-L occurred, a gradual increase in the acenaphthene fraction over time would be expected. Indeed, the fractional distribution of PAH-L compounds showed slight variation throughout the experiment, with acenaphthene increasing from 71 % at the start to 87 % by the end, while naphthalene decreased from 26 % to 12 % (Fig. 3), indicating that some degradation might have occurred. Nevertheless, since the soil was treated applying EO, it was not possible to differentiate between the electrochemical and biological degradation processes, making it unclear which was causing this shift. The concentration of acenaphthylene remained the most stable, accounting for only 1–3 % of the PAH-L fraction, and was negligible compared to the other PAH-L compounds (Fig. 3).

Among the PAH-M compounds, fluoranthene and pyrene were predominant, together accounting for 99 % of the PAH-M fraction and 74 % of the total PAH16 at the start of the experiment. By the end, these two compounds, which are among the most toxic PAHs in creosote (U.S. Environmental Protection Agency, 2008), showed the greatest reductions, decreasing by 92 % and 93 %, respectively. This suggests that the soil treatment had the most significant impact on the PAH-M fraction, particularly on fluoranthene and pyrene, but PAH-H in the soil solution also showed a visible decline, decreasing from 384 µg/L to 143

<figure><img src="image_6.png" /><figcaption>**Fig. 3.** Fractional distribution of PAH-L constituents (naphthalene, acenaphthylene, and acenaphthene) in soil solution over the course of the experiment. S1-first season, S2-second season.</figcaption></figure>
µg/L over the course of the experiment (i.e. 63 % decrease).

In a study of creosote-contaminated soil by Smulek et al. (2020) a 40–90 % reduction in fluoranthene and pyrene was observed through biological degradation. Similar degradation rates were reported by Eriksson et al. (2003) in soils from arctic regions at 20 °C. However, the degradation rates dropped significantly when the temperature decreased to 7 °C. During the first year of our experiment, outdoor temperatures dropped from 14.8 °C to 5.8 °C by the end of the season, while PAH concentrations in the soil solution reached their lowest levels during the final sampling in October of the first season. In the second season, despite temperatures rising from 14 °C to 22 °C at the beginning of the sampling, the PAH concentrations in the solution remained similar to those observed in September of the first season (Fig. 1). The increase in temperature could contribute to the dissolution of PAH and it has been reported that the solubility of PAH such as naphthalene, phenanthrene, anthracene, pyrene, and benzo(a)pyrene in water can increase two to three times when the temperature rises from 12 °C to 30 °C (Schwarz, 1977). Nevertheless, higher temperatures during the second season did not appear to enhance PAH degradation. Previous observations by Eriksson et al. (2003) have shown that the composition of enriched microbial communities in creosote-contaminated northern soils was more influenced by the presence of oxygen than by temperature. Thus, it is important to ensure that wet conditions do not persist, as this could limit oxygen availability for aerobic microorganisms, potentially hindering PAH biodegradation in soil. However, optimal electricity propagation occurs in water-saturated soil, making sufficient water levels preferable for electrochemical oxidation. This might be the reason for the observed drop in PAH levels in soil solution in October, when the soil was most saturated, despite the decrease in redox potential. This creates a need for balance between maintaining moisture for effective electrochemical treatment and avoiding excessive saturation that might limit activity of aerobic PAH degraders. Under anaerobic conditions, the potential for PAH biodegradation is highly limited (Sharak Genthner et al., 1997; Haritash and Kaushik, 2009).

### 3.4. Impact of electrooxidation on PAH in soil

In addition to the dominant PAHs identified in the soil solution, the soil also contained significant amounts of phenanthrene (PAH-M), benzo

(k)fluoranthene, and benzo(a)pyrene (both PAH-H), which all together accounted for 69–92 % of the total PAH concentration in the soil (Table S2).

The presence of free-phase creosote oil contamination in the soil made the determination of PAH16 in the soil samples very unreliable. That is, some soil particles were coated with oil, while others were not. This made the proper soil homogenization and representative sampling nearly impossible despite that a cement blender was used to mix and homogenise the soil before placing it into the boxes. Bulk soil samples that were taken directly after mixing and analysed for PAH16 (Table 1) showed considerably different results from the samples taken from treated and untreated soil at the end of each season (Table 2). At the end the first season (C₁) and the second season (C₂) soils were taken from the boxes at different locations as described earlier. The analysis results indicate an extensive variation in PAH16 concentrations, differing by an order of magnitude between the sampling occasions and points (Table 2). This needs to be considered when interpreting the results and discussing the practical applicability of the technology. An enrichment of PAH was observed at one of the electrodes (E2) in peat compared to the other sampling points, but not in sand (Table 2). Although a movement of PAH through electroosmotic flow can occur (Alcántara et al., 2012; Han et al., 2021), the polarity reversal should have prevented any accumulation of PAH at one of the electrodes. Therefore, such differences in PAH concentrations in soil, also observed between treated and untreated samples, could barely be caused by the treatment and can only be explained by the initial heterogeneity of the PAH distribution in soil.

It is therefore difficult to conclude whether the EO treatment caused a notable decrease in PAH16 concentrations throughout the entire treated soil volume, in contrast to the clear reduction observed in the soil solution.

Nevertheless, when comparing only the results from treated soil collected at the end of the second season to those from the end of the first season, PAH16 concentrations significantly decreased at 10 out of the 12 sampling points in both sand (on average 84 ± 13 % reduction) and peat (69 ± 28 % reduction) (Fig. 4). It is important to note, that in difference to the samples taken on bulk soil, these samples from the end of C₁ and C₂ were taken from the very same sampling points and is hence a more accurate measurement of the treatment success. This suggests that PAH

<table><caption>**Table 2**
Concentration of PAH16 (mg kg⁻¹) in the untreated soil samples, compared to the soil collected after the first and the second seasons.</caption><thead><tr><th rowspan="2">Sample</th><th>PAH 16<br>mg kg<sup>-1</sup></th><th>Average<br>treated mg kg<sup>-1</sup></th><th>Change<br>%</th><th>PAH 16<br>mg kg<sup>-1</sup></th><th>Average<br>treated mg kg<sup>-1</sup></th><th>Change<br>%</th></tr><tr><th colspan="2">Season 1</th><th colspan="3">Season 2</th></tr></thead><tbody><tr><td>Untreated Peat</td><td>Box 1<br>20,400</td><td></td><td></td><td>Box 1<br>4090</td><td></td><td></td></tr><tr><td>Treated Peat centre</td><td>6050</td><td></td><td></td><td>2240</td><td></td><td></td></tr><tr><td>Treated Peat electrode 1</td><td>5990</td><td>28,213</td><td>+38</td><td>4160</td><td>2267</td><td>-45</td></tr><tr><td>Treated Peat electrode 2</td><td>72,600</td><td></td><td></td><td>402</td><td></td><td></td></tr><tr><td>Untreated Peat</td><td>Box 2<br>21,800</td><td></td><td></td><td>Box 2<br>16,700</td><td></td><td></td></tr><tr><td>Treated Peat centre</td><td>5060</td><td></td><td></td><td>11,500</td><td></td><td></td></tr><tr><td>Treated Peat electrode 1</td><td>2680</td><td>5430</td><td>-75</td><td>1040</td><td>4393</td><td>-74</td></tr><tr><td>Treated Peat electrode 2</td><td>8550</td><td></td><td></td><td>638</td><td></td><td></td></tr><tr><td>Untreated Sand</td><td>Box 1<br>5340</td><td></td><td></td><td>Box 1<br>206</td><td></td><td></td></tr><tr><td>Treated Sand centre</td><td>2200</td><td></td><td></td><td>751</td><td></td><td></td></tr><tr><td>Treated Sand electrode 1</td><td>242</td><td>1201</td><td>-78</td><td>3110</td><td>1380</td><td>+570</td></tr><tr><td>Treated Sand electrode 2</td><td>1160</td><td></td><td></td><td>278</td><td></td><td></td></tr><tr><td>Untreated Sand</td><td>Box 2<br>180</td><td></td><td></td><td>Box 2<br>32</td><td></td><td></td></tr><tr><td>Treated Sand centre</td><td>1900</td><td></td><td></td><td>70</td><td></td><td></td></tr><tr><td>Treated Sand electrode 1</td><td>2590</td><td>1669</td><td>+827</td><td>145</td><td>93</td><td>+194</td></tr><tr><td>Treated Sand electrode 2</td><td>517</td><td></td><td></td><td>64</td><td></td><td></td></tr></tbody></table>
<figure><img src="image_7.png" /><figcaption>**Fig. 4.** Concentration PAH16 in ten sampling points of A) sand and B) peat after the first season (S1) and after the second season (S2) in the centre of the two boxes (C1, C2) and at the electrodes (E1-1, E1-2 and E2-1, E2-2).</figcaption></figure>

degradation may have occurred, and the decrease in PAH16 concen-
tration over time suggests that even longer treatment time is needed to
obtain reduction in all locations, particularly for soils with such high
PAH16 concentrations as those observed in this study.

Contrary to the soil solution, the greatest reduction in soil at those
ten points was observed for PAH-L, followed by PAH-M, and lastly PAH-
H, with this trend consistent in both sand and peat. In sand, PAH-L
concentrations decreased by 92 ± 8 %, PAH-M by 85 ± 14 %, and
PAH-H by 70 ± 17 %. In peat, the reductions were on average less
pronounced, with PAH-L decreasing by 87 ± 12 %, PAH-M by 70 ± 30
%, and PAH-H by 57 ± 32 %. This difference is most likely due to the
high organic matter (OM) content in peat, which has been previously
shown to hinder PAH removal in EO-treated soil (Saberi et al., 2018).
These results suggest that longer remediation times are required in soils
with high OM content compared to sandy soils.

Various Fe compounds have been found to improve microbial
degradation of petroleum hydrocarbons, for example by serving Fe(III)
as an electron acceptor in microbial hydrocarbon degradation (Castro
et al., 2022). This suggests that corrosion products from Fe electrodes
could have a positive effect on PAH degradation. In the present study,
the treatment lead to an increase in the fractions of active Fe species
(Johansson et al., 2024), which may support both chemical and bio-
logical pathways for PAH degradation. Also, corroding Fe provides new
surfaces for As sorption and its immobilisation, leading to the reduced
bioavailability of As in soil (Kumpiene et al., 2023). By this, the soil
toxicity for bacteria would also decrease. Using an alternating polarity,
contaminants are confined within the treated area, which can prolong
the time available for contaminant desorption and degradation (Huang
et al., 2012; Mena et al., 2016). In addition, alternating polarity prevents
soil acidification at the anode and alkalinization at the cathode - con-
ditions that are often harmful to soil microorganisms and commonly
occur during traditional EO treatments (Chilingar et al., 1997; Saichek
and Reddy, 2005).

3.5. *Impact of electrooxidation on soil microorganisms*

The degradation of the PAH compounds observed in the soil solution
is believed to result not only from the electrooxidation, but also from the

activity of the microbial population in soil (Li et al., 2012). Moreover, it
has been shown that combining EO with other treatments, such as
bioremediation can enhance the remediation efficiency, shorten the
remediation time and lower the energy cost (Huang et al., 2012). All the
physicochemical phenomena affecting the soil have an influence on the
diversity and the structure of the microbial community, where only the
microorganisms able to tolerate or degrade the PAHs will prevail. Under
these circumstances, time was the most important factor reshaping the
microbial population. In Fig. S4, the principal coordinate analysis
(PCoA) shows that time is the factor driving shifts in microbial structure,
regardless of the soil type, weather it is sand or peat. Given the presence
of pollutants in all samples, bacteria capable of degrading or tolerate
these contaminants naturally become more dominant over time. The
different bacterial taxa able to adapt to the new environment, i.e.
degrading pollutants, or tolerate the electrochemical changes, survive as
they acclimate to these conditions.

To provide a clearer understanding of the changes in the different
taxa during the process, Fig. 5 illustrates the phenomena through
various boxplots. Overall, no significant changes in the DNA concentration
over time could be observed for any of taxa. However, some differences
are visible, depending on the soil type. The *Firmicutes* and the *Chloroflexi*
phyla were slightly more prevalent in the sand samples, regardless of the
season, which contained lower concentrations of PAH compared to peat.
*Desulfobacterota*, which is known for using sulfate as terminal electron
acceptor in anaerobic respiration and for its ability to perform exocel-
lular electron transfer also showed distinct patterns with a slight
decrease observed in peat over time.

The *Gammaproteobacteria* were the most prevalent phylum in all
samples, with a slightly higher average concentration in samples from
the second season (Fig. 5). The dominance of *Gammaproteobacteria* is
characteristic for contaminated soils. Niepceron et al. (2013) suggested
using this bacterial class as a bioindicator when assessing the impact of
PAH contamination on bacterial communities in soil due to adaptability,
PAH-degrading capacity and prevalence of *Gammaproteobacteria* in
contaminated environments. Furthermore, *Gammaproteobacteria* was
observed to be the dominant class in soil highly contaminated with
chromium, while no members of this class were present in the control
(non-contaminated) soil (Thompson and Wickham, 2018), thus showing
<figure><img src="image_8.png" /><figcaption>**Fig. 5.** Variation of the different taxa concentration (µg of DNA mL⁻¹) across the different sampling points (A1, start of the first season; A2, end of first season, B1, start of the second season; B2, end of second season).</figcaption></figure>

the resistance and adaptability of this bacteria class to a broad range of soil contaminants. The soil in this study contained high concentrations of As (150 mg kg⁻¹ in sand and 438 mg kg⁻¹ in peat) (Johansson et al., 2024). Over time, the concentrations of dissolved As decreased due to the soil treatment, with the most significant reduction observed in the sandy soil (by 50–54 %). This change may explain the shift in microbial abundance between the two soil types. Initially, *Gammaproteobacteria* were more abundant in the peat, but by the end of the experiment, their abundance slightly increased in the sand, which could indicate that *Gammaproteobacteria* have some sensitivity. Kurm et al. (2017) has demonstrated that *Gammaproteobacteria* isolated from soil and further cultivated in laboratory were growing fastest and were able to utilize a higher number of substrates than strains belonging to other bacterial classes (*Actinobacteria*, *Bacilli*, *Betaproteobacteria* and *Alphaproteobacteria*). Despite concerns over antibiotic resistance (Broszat et al., 2014), species such as *Pseudomonas* and *Stenotrophomonas* within *Gammaproteobacteria* class have been associated with plant health in the rhizosphere (Köberl et al., 2017), indicating their complex role in both soil health and contaminated soil environments.

The ASVs were normalized using the DeSeq2 package in R, with their abundance displayed individually for each sample. The 64 most abundant ASVs are visualized in a heatmap (Fig. 6), where the ASVs are clustered based on Euclidean distance according to their abundance patterns. Four distinct clusters emerged. The first cluster includes ASVs whose abundance gradually increases over time. Interestingly, four out of the seven samples in this cluster belong to the *Sulfuritalia* species, a member of the *Rhodocyclaceae* family. The second cluster is more diverse, comprising ASVs that tend to diminish over time. The third cluster is characterized by ASVs that disappear over the course of the

experiment, likely reflecting the toxic effects of PAHs. In this cluster, eight ASVs remain unidentified, while the majority of the identified ASVs belong to the *Alphaproteobacteria* class.

The physicochemical changes occurring in the soil are mirrored in the porewater composition. The detailed information on changes in As and Fe concentrations in porewater are given in Johansson et al. (2024). Here, the parameters were correlated with the 20 most abundant ASVs (Fig. S5). Four positive correlation groups are evident, with the 20 ASVs grouped into three distinct clusters. Arsenic and Fe concentrations, pH and EC were grouped together, indicating their decreasing trends over time are similar. Notably, these factors correlated negatively with the ASVs group that increased in abundance over time. The same group also included the redox potential, indicating that as redox potential increased, the concentrations of dissolved As and Fe decreased, which is the expected behaviour of redox sensitive elements in soil.

## 4. Conclusions

In this study, an attempt was made to degrade PAH16 *in situ* using electrochemical oxidation (EO) in two 1 m³ outdoor boxes. Over the course of the two-season experiment, PAH16 concentrations in the soil solution decreased by an average of 82 %. This reduction was primarily observed in the medium and high molecular weight fractions, suggesting that EO using low-voltage pulsing current with alternating polarity can effectively reduce the concentrations of more recalcitrant PAH compounds. The most significant reduction was observed for fluoranthene and pyrene (92 % and 93 %, respectively), which are among the most toxic PAHs in creosote. The largely unchanged concentration of low molecular weight PAHs in the solution was likely due to replenishment
<figure><img src="image_9.png" /><figcaption>**Fig. 6.** Heatmap with the 50 most abundant ASVs according to differential expression analysis. Values were normalized using the variance stabilization method, and used to cluster the ASVs. Samples are vertically organized, and ASVs are displayed horizontally.</figcaption></figure>

from the PAHs sorbed to the soil, a process potentially triggered by the EO treatment. During the same time, concentrations of PAH16 in soil samples decreased in 10 out of the 12 sampling points, with a greater average reduction in sand (84 % decrease) compared to peat (69 % decrease).

The microbiological tests showed that the *Gammaproteobacteria* were the most prevalent phyla in all soil samples throughout the treatment, confirming the ability of this class to thrive in highly contaminated soil environments. Overall, the abundance of the identified bacterial taxa did not change significantly during the course of the treatment.

Overall, this study suggests that EO treatment can be a suitable method for *in situ* remediation of PAH16 in creosote-contaminated soil, but longer remediation times are necessary to achieve effective PAH degradation throughout the entire volume of treated soil, particularly in soils with high organic matter (OM) content. Moreover, due to the heterogeneity of soil samples, especially when containing free phase creosote, more studies are needed to confirm these results prior industrial scale implementation.

### CRediT authorship contribution statement

**Kim Johansson:** Writing – original draft, Visualization, Validation, Resources, Methodology, Investigation, Formal analysis, Conceptualization. **Azriel Ruiz-Valencia:** Writing – review & editing, Visualization, Validation, Resources, Methodology, Investigation, Conceptualization. **Ivan Carabante:** Writing – review & editing,

Visualization, Validation, Supervision, Resources, Methodology, Investigation, Formal analysis, Conceptualization. **Timothy M. Vogel:** Writing – review & editing, Visualization, Validation, Supervision, Resources, Project administration, Methodology, Investigation, Formal analysis, Conceptualization. **Erkki Lindberg:** Visualization, Validation, Resources, Methodology, Conceptualization. **Jurate Kumpiene:** Writing – review & editing, Visualization, Validation, Supervision, Resources, Project administration, Methodology, Investigation, Funding acquisition, Formal analysis, Conceptualization.

### 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.

### Acknowledgements

This project has received funding from the European Union's Horizon 2020 research and innovation program under grant agreement N°965945 and the Swedish Research Council for Sustainable Development FORMAS, project no. 2022-00864. All EiCLaR partners are acknowledged for contributing discussions. Ekogrid are acknowledged for contributing with the power unit. Dr Mariusz Gusiatin and Tanise Yang are acknowledged for their help with sampling and monitoring.
# Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.jenvman.2025.125978.

## Data availability

Data will be made available on request.

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