# Ultrasound-assisted removal of organochlorine pesticide chlordecone from aqueous matrices

Maxime Cochennec, Deborah E Martin, Benjamin Laulier, Alexandre

Baudouard, Pascal Tierce, E.D. van Hullebusch, Julien Parinet, Stéfan

Colombano

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Maxime Cochennec, Deborah E Martin, Benjamin Laulier, Alexandre Baudouard, Pascal Tierce, et al.. Ultrasound-assisted removal of organochlorine pesticide chlordecone from aqueous matrices. Environmental Technology and Innovation, 2025, 40, pp.104563. 10.1016/j.eti.2025.104563. hal-05333651

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# Ultrasound-assisted removal of organochlorine pesticide chlordecone from aqueous matrices

Maxime Cochennec<sup>a, *, †</sup>, Deborah E. Martin<sup>d</sup>, Benjamin Laulier<sup>c</sup>,

Alexandre Baudouard<sup>c</sup>, Pascal Tierce<sup>c</sup>, Eric D. van Hullebusch<sup>b</sup>, Julien Parinet<sup>d</sup>,

Stéfan Colombano<sup>a</sup>

<sup>a</sup> BRGM, Orléans F-45060, France

<sup>b</sup> Université Paris Cité, Institut de Physique du Globe de Paris, CNRS, Paris F-75005, France

<sup>c</sup> SinapTec, 7 avenue Pierre et Marie Curie, Lezennes 59260, France

<sup>d</sup> ANSES, Laboratory of Food Safety, Maison-Alfort F-94701, France

## ARTICLE INFO

### Keywords:

- Chlordecone

- Pesticide

- Sonochemistry

- Ultrasound

- Degradation

## ABSTRACT

Chlordecone (CLD) is a persistent organochlorine pesticide that causes adverse human health effects and remains widespread in certain regions, such as the French West Indies. Despite its hydrophobicity, CLD is detected in surface and groundwater, sometimes at concentrations exceeding regulatory standards. Interestingly, this same hydrophobicity makes chlordecone a suitable candidate for removal via sonolysis. Sonolysis, which is based on the creation and collapse of cavitation bubbles in a liquid phase, has been proven effective in removing organochlorine pesticides. This study is the first to demonstrate that sonolysis can effectively remove CLD from both deionized spiked water and natural surface water at environmentally relevant CLD concentrations (0.4–4 µg L⁻¹). Approximately 74 % of CLD is degraded within 120 min, regardless of whether the water matrix is synthetic or natural. Removal proceeds slightly faster in surface water samples, likely due to the suspended solids on which CLD tends to bind, which catalyze the formation and growth of cavitation bubbles. This highlights that the likely mechanism for removal is the thermal breakdown of CLD at the bubble interface rather than oxidation in the bulk. The performance, evaluated through the Electrical Energy per Order (EEO) figure of merit, shows that sonolysis outperforms photolysis (444 kW h m⁻³ vs 1600 kW h m⁻³), although the CLD concentration differs. The EEO decreases – hence, energy efficiency increases with increasing CLD concentration. Future studies should focus on determining the maximum concentration of CLD before reactive sites become saturated – a prerequisite for scaling up sonolysis.

## 1. Introduction

Chlordecone (CLD; C₁₀Cl₁₀O; CAS number 143–50–0) is a persistent organochlorine pesticide that has been linked to adverse human health effects (Multigner et al., 2016), including, among others, an increased risk of prostate cancer (Multigner et al., 2010), impaired fine motor skills (Boucher et al., 2013), increased risk of premature birth (Kadhel et al., 2014) and endocrine disruption (Cordier et al., 2015). CLD was used, not exclusively, on banana plantations between 1972 and 1993 to control the banana weevil in

E-mail address: m.cochennec@brgm.fr (M. Cochennec).

https://doi.org/10.1016/j.eti.2025.104563

Received 25 June 2025; Received in revised form 27 September 2025; Accepted 9 October 2025

Available online 10 October 2025

2352-1864/© 2025 The Authors. Published by Elsevier B.V. This is an open access article under the CC BY license (http://creativecommons.org/licenses/by/4.0/).

<footnote>* Corresponding author.</footnote>
the French West Indies (Guadeloupe, Martinique). Since then, contamination of the environment in these territories has been a significant health, environmental, and ecological crisis (Levillain et al., 2012). Despite the global ban of CLD, sometimes marketed as Kepone®, by the Stockholm Convention in 2009 (Madaj et al., 2018), CLD can still be found in soils (Rochette et al., 2020), groundwater (Crabit et al., 2016), surface water (Crabit et al., 2016; Voltz et al., 2023), sediments, or animal organisms (Coat et al., 2006). This is because CLD is remarkably persistent in the environment due to its high thermodynamic stability and is sometimes considered nonbiodegradable in the environment (Dolfing et al., 2012). Field observations have recently challenged this view, as a CLD half-life of 5 years is more effective in reproducing the field observations according to the authors' model (Comte et al., 2022). However, these findings are debated (Saaidi et al., 2023). In this regard, while theoretical reasoning allows for the biotransformation of CLD (Dolfing et al., 2012), and anaerobic biodegradation has been observed multiple times in the laboratory (Lomheim et al., 2021; Martin et al., 2023; Hellal et al., 2021; Chaussonnerie et al., 2016), the potential environmental biotic degradation pathway remains unknown.

CLD is also poorly soluble in water, ranging from 2 mg L⁻¹ to 3 mg L⁻¹ (Faroon et al., 1995), and tends to accumulate at soil particle surfaces or in fatty/oily phases. Therefore, CLD transfer through animal products and crop consumption is one of the most significant exposure pathways (Dromard et al., 2022, 2018; Jondreville et al., 2014). Consequently, targeted soil remediation actions have been advocated and developed for hotspots. Soil remediation techniques include well-studied techniques such as reductive dechlorination with nano-zero-valent iron amendment (Mouvet et al., 2017, 2020; Belghit et al., 2015; Dictor et al., 2011) or anaerobic biodegradation (Lomheim et al., 2021; Chaussonnerie et al., 2016). New techniques such as microwave heating (Cochennec et al., 2024) or aerobic biodegradation (Fernández-Bayo et al., 2013) have also been developed. However, soils are not the only important compartment, since the “low” solubility of CLD is nonetheless approximately 20,000 times higher than the value required for drinking water under French law (0.1 µg L⁻¹). Furthermore, despite its high K₀c value between 2000–2500 L kg⁻¹ (Cabidoche et al., 2009), and high octanol-water partition coefficient, log(K₀w) = 4.5 (Coat et al., 2006), groundwater has also been shown to be a significant transfer path from the hotspots (i.e., soils) to other environmental compartments (surface water, sediments, etc.), both by the dissolved CLD or by the particle-bound CLD fraction (Crabit et al., 2016; Faroon et al., 1995; Bizeul et al., 2025). Consequently, adequate treatment of the contaminated water is also needed, besides soil remediation (Della Rossa et al., 2017).

The removal of CLD from an aqueous matrix can be achieved by adsorption on a substrate, e.g., activated carbon (Durimel et al., 2013), activated carbon coated with cyclodextrins (Rana et al., 2016), hemicryptophane supramolecule (Della-Negra et al., 2024), or activated carbon doped with iron oxide nanoparticles (Rana et al., 2017). Degradation methods are preferred as they don't produce a residual phase that must be managed, although the formation of by-products should also be considered. In this regard, the potential degradation by-products of CLD are known to be more mobile, as the solubility tends to decrease with the number of chlorine atoms (based on SPARC online calculator (Dolfing et al., 2012; Hilal et al., 2004), but their toxicity is not well understood. A recent study based on *Hydra vulgaris* bioassay has shown that the toxicity of CLD mixed with various CLD byproducts from dechlorination is similar to that of CLD alone, indicating that CLD byproducts do not pose a significant risk to aquatic environments in terms of toxicity (Moreau et al., 2022). The chemical reduction of CLD in water matrices has been studied, for example, using vitamin B12 as a catalyst (Schrauzer and Katz, 1978). This technique has recently been combined with sequestration by activated carbon, followed by ethanol/methanol desorption and subsequent reductive dechlorination mediated by vitamin B12 (Ranguin et al., 2017). The degradation of CLD from an aqueous matrix by oxidation includes methods such as photolysis, which has shown good degradation efficiency (Cruz-González et al., 2018). In contrast, the same authors compared it with other advanced oxidation processes, such as ozonation or Fenton oxidation, but found no or minor efficiency. This work is also noteworthy because the authors calculated each technique's electrical energy per order (EEO), a standard comparison parameter for advanced oxidation processes (Bolton et al., 1996). The EEO is defined as

$$EEO = \frac{P \times t}{V\ln(C_0/C_t)} \qquad (1)$$

where P (kW) is the power, t (h) is the reaction time, V (m³) is the volume of solution, C₀ is the initial concentration, and Cₜ is the concentration at time t. The authors found values 1600 and 1890 kW h m⁻³ for photolysis and photo-Fenton CLD degradation, respectively. Eq. 1 can also be expressed as

$$EEO = \frac{P}{V} \qquad (2)$$

where k is the degradation rate constant obtained by fitting a first-order kinetic model to the experimental data, as log (C/C₀) = -kt.

The sonolysis method is based on the propagation of ultrasonic acoustic waves into a liquid bath, generating vapor cavitation bubbles that live for a very short time (microseconds) and then collapse. Bubbles collapsing locally generate very high temperatures (> 5000 °C) and pressures (> 1000 bars), which lead to the thermal decomposition of organic compounds present near the bubbles or at the liquid/gas interface, and as such, can be used for the remediation of aqueous matrices from organic pollutants (de Andrade et al., 2021; Pirsaheb and Moradi, 2020), besides other usage such as extraction (Huang et al., 2022, 2025) and synthesis (Sabirov et al., 2025; Kinzyabaeva and Sh, 2020). The exact degradation mechanisms and removal efficiency depend on the molecule to degrade and the operating parameters. These include, primarily, the frequency [Hz], typically ranging between low (12 kHz) to high (1000 kHz) frequency, the power density [W L⁻¹], the reaction time [min], and the initial pollutant concentration [mg L⁻¹] (de Andrade et al., 2021; Pirsaheb and Moradi, 2020). It is sometimes believed that low-frequency treatment primarily involves the creation of hydroxyl
radicals following a reaction such as H₂O + ))US → H⁺ + ·OH, which then causes oxidative degradation. In contrast, at high frequencies, it is sometimes believed that the thermal breakdown of the molecule is the most significant phenomenon. However, this is a simplified picture, as the two mechanisms largely happen simultaneously but at different zones (e.g., thermal decomposition in the bubble core and oxidation due to radicals' production near the bubble interface) (de Andrade et al., 2021). More generally, the efficiency of sonolysis in degrading organic pollutants is considered maximum in the range of 200–600 kHz (Petrier and Francony, 1997). Because thermal breakdown is not the only mechanism for degradation, sonolysis is sometimes referred to as an advanced oxidation process (Malakootian et al., 2020; Mishra et al., 2017).

Sonolysis has been applied to various organochlorine pesticides, but not to CLD, despite being considered one of the most effective methods for treating aqueous solutions contaminated by organochlorine pesticides (da Silva Júnior et al., 2024). Researchers have demonstrated that sonolysis is capable of degrading up to 70 % of a cocktail of organochlorine pesticides (22 compounds at 20 or 40 µg L⁻¹ each, including heptachlor (C₁₀H₅Cl₇), lindane (C₆H₆Cl₆), or decachlorobiphenyl (C₁₂Cl₁₀)) (Kida et al., 2018). The sonolysis tests were performed using a low-frequency probe (20 kHz, nominal power 200 W) at 20 % amplitude, and the authors assessed the process efficiency by dosing the total organic content. Sonolysis itself appeared more effective than the hybrid process of sonolysis – oxidation by H₂O₂. Moreover, GC-MS chromatograms showed that sonolysis failed to degrade completely organochlorine pesticides such as lindane or heptachlor (Kida et al., 2018). Sonolysis has also proved effective in degrading DDT, a highly persistent organochlorine pesticide (Thangavadivel and Megharaj, 2009). A 90-min treatment degraded up to 90 % of the pesticide initially spiked in water (1200 µg L⁻¹) using a probe at a frequency of 1600 kHz and with a power density of 150 W L⁻¹. The authors also measured chloride concentration and found that DDT degradation, while forming intermediate by-products, tended to approach complete mineralization over time. Sonolysis has also successfully degraded dicofol, an organochlorine pesticide considered a persistent organic pollutant (Deabrata and Sivakumar, 2018). Using a 20 kHz probe, the authors found a maximum removal yield of 70 % at a power density of 3750 W L⁻¹ (based on the probe's nominal power). The authors also observed some degradation by-products (Deabrata and Sivakumar, 2018). Additionally, two works report the degradation of pentachlorophenol (C₆HCl₅O) by sonolysis. The first one employed a 20 kHz probe or a 500 kHz orthoreactor (Weavers et al., 2000). The authors found that the degradation kinetic was similar for the low-frequency probe and the high-frequency orthoreactor, contrasting with the ideal frequency range, which is supposed to be between 200 and 600 kHz. This can be explained by the difference in power density between the two tests, as obtained by calorimetric measurements, 75.5 W L⁻¹ for the high-frequency reactor and 1340 W L⁻¹ for the low-frequency probe. The second study on pentachlorophenol degradation highlights that a dual-frequency system enhances degradation, with the best results obtained at 20/530 kHz (Wang et al., 2006). The authors used a titanium probe operating at 20 kHz and an ultrasonic emitter based on a ceramic disk for the higher frequency. They fitted the results to a pseudo-first-order kinetics reaction, from which the reaction rate coefficient was determined to be 0.033 min⁻¹ in the best tested conditions.

Sonolysis has not been applied to CLD degradation despite its proven effectiveness in degrading other organic pollutants, including organochlorine pesticides. It should be noted that CLD is a particularly stable organochlorine pesticide due to its molecular cage-like structure and is more difficult to degrade than many other organochlorine pesticides. For example, whereas CLD environmental half-life is estimated in decades or more, DDT half-life is about 2–15 year, Dicofol half-life is about 60 days, and lindane half-life is 15 months (Jayaraj et al., 2016). This study aims to assess the efficacy of sonolysis, without the use of additional reagents and at ambient pressure/temperature conditions, in degrading CLD. The matrix effect (treatment of contaminated surface water samples and spiked deionized water) and the impact of initial concentration (1–10 times the concentration in contaminated surface water) are investigated. The energy budget of the technique is discussed and compared to other advanced oxidation processes already applied and reported in the literature using the EEO. By doing so, application to water treatment or concentrated leachate can be envisaged with a complete understanding of the results that can be obtained under controlled laboratory conditions.

## 2. Material and methods

### 2.1. Chemicals

Chlordecone (CLD, C₁₀Cl₁₀O₂H₂) and ¹³C₁₀-CLD (purity >98 %) were supplied by Azur Isotopes, Marseille, France. A Milli-Q system obtained from Merck Millipore (Saint-Quentin-en-Yvelines, France) was used to generate ultrapure water (18.2 MΩ.cm). Formic acid, acetonitrile (MeCN), and dichloromethane (DCM) (LC-MS-grade) used for the extraction and/or analyses were obtained from Fisher Scientific (Illkirch, France).

### 2.2. Environmental samples

The surface water was sampled without filtration in the Pérou river from the OPALE ("Observatoire des Pollutions Agricoles aux Antilles") Observatory. The Pérou River watershed presents a very high risk of CLD contamination, as it is home to numerous banana plantations (Sabatier et al., 2021). The TSS analysis was carried out in accordance with standard NF EN 872. This involves filtering a volume of water through a glass fiber membrane, then drying it at 105°C and weighing the filter to determine the mass of suspended matter.

The result is expressed in mg L⁻¹, corresponding to the difference in mass of the filter before and after filtration relative to the sample volume.
## 2.3. Sonolysis protocol

The sonolysis experiments were conducted on deionized water (DI water) spiked with CLD and contaminated surface waters sampled in Guadeloupe. The main experimental setup consisted of an ultrasonic generator operating at 500 kHz with a nominal power of 100 W, a transducer, and a stainless-steel jacketed reactor with a total capacity of 500 mL. The whole sonolysis system is thus more akin to a “bath-type” system than a probe and was developed and provided by SinapTec®. The setup was more advanced than a mere ultrasonic bath, as the equipment was automatically regulated to match the ideal frequency around 500 kHz, thus ensuring an optimal power delivery. Auxiliary equipment involved a chiller bath to regulate the reactor temperature at 20°C through the jacket, a PT100 temperature probe, and a bottle of compressed air to cool down the transducer.

The sonolysis procedure was as follows: First, 315 ± 1 mL of contaminated aqueous solution was poured into the reactor. The solution was either the sampled surface water or deionized water spiked with CLD at various concentrations. The chiller bath was then set to maintain a temperature close to 25°C, the target temperature. A first 15 mL sample was taken immediately before starting the generator. Then, the generator starts, and the sonolysis lasts 2 h. 15 mL samples were withdrawn at times 15, 30, 60, and 120 min. Two experiments were conducted under a given set of initial conditions to check the reproducibility of the experiments. The power density was maintained at 333 W L⁻¹ by proportionally decreasing the generator power per the volume of liquid remaining in the reactor. The samples were stored in 20 mL glass tubes and refrigerated until the chemical analysis was performed.

# 3. Chemical analysis

## 3.1. Liquid-liquid extraction

Extractions of the water samples were performed with a liquid/liquid extraction; 10 mL of water samples were transferred to a 50 mL centrifuge tube. Then, 20 µL of a solution of ¹³C¹⁰-CLD at 250 ng mL⁻¹ were added (final concentration of 0.5 µg L⁻¹), and samples were vortexed for 1 min. Ten (10) mL of DCM were poured into the centrifuge tube and vortexed for 1 min. The organic phases (DCM), in the bottom of the centrifuge tube were recovered using a Pasteur pipette, evaporated to dryness, and reconstituted in 1 mL of MeCN and deposited in a 2 mL amber vial prior to LC-MS/MS analysis with isotopic dilution and bracketing.

## 3.2. HPLC-MS/MS analysis

The current HPLC-MS/MS analysis method is described by (Saint-Hilaire et al., 2018). Analyst 1.5.1 software (AB Sciex, Les Ulis, France) was used for equipment control and data acquisition. Following the SANTE guidelines (Sante/11312/2021), two consecutive single reaction monitoring (SRM) runs were realized for CLD (506.7/426.7 and 508.7/428.7). Selected identification criteria were retention time deviation and the difference between the two SRM ratios of less than 0.1 min and 30 %, respectively. Instrument response linearity was considered between 0.05 and 220 ng mL⁻¹ of CLD in ACN using a constant concentration of ¹³C¹⁰-CLD (0.5 µg L⁻¹). The limit of quantification (LOQ) was set to 0.05 µg L⁻¹ for CLD.

# 4. Results

## 4.1. CLD quantification

The initial concentration in CLD for each experiment, as well as the aqueous matrix type, is given in Table 1. Surface water showed an initial concentration in the range of 0.30–0.36 µg L⁻¹ (Experiments 1 & 2). Consequently, deionized water was spiked at a similar level and shows an initial CLD concentration of 0.41 µg L⁻¹. Additionally, deionized water prepared with more CLD showed an initial concentration ranging from 3.45 to 3.99 µg L⁻¹. In comparison, the study on photolysis and photo-Fenton oxidation used demineralized water spiked with an initial CLD concentration of 1000 µg L⁻¹ (Cruz-González et al., 2018). For example, CLD concentrations of up to 2.4 µg L⁻¹ have been reported in surface water (river) (Crabit et al., 2016). In the same work, CLD concentrations of 0.65 µg L⁻¹ and 3.24 µg L⁻¹ were reported for groundwater in Nitisol and Andosol conditions, respectively. Consequently, the matrices investigated in this study correspond, in terms of CLD concentration, to the average and upper range of environmental CLD concentration in the French West Indies.

<table><caption>**Table 1**
Initial concentration in CLD and aqueous matrix type for each experiment.</caption><thead><tr><td>Exp #</td><td>Matrix</td><td>Initial CLD concentration (µg L<sup>-1</sup>)</td></tr></thead><tbody><tr><td>1</td><td>Surface water</td><td>0.30</td></tr><tr><td>2</td><td>Surface water</td><td>0.36</td></tr><tr><td>3</td><td>Deionized water</td><td>3.99</td></tr><tr><td>4</td><td>Deionized water</td><td>0.41</td></tr><tr><td>5</td><td>Deionized water</td><td>3.45</td></tr><tr><td>6</td><td>Deionized water</td><td>0.41</td></tr></tbody></table>
## 4.2. Aqueous matrix effect

The CLD removal after 2 h reached 75 % and 73 % for the surface water samples, and 72 % and 73 % for the deionized water, at similar CLD concentration. The reaction time was sufficient to decrease the CLD concentration in the surface water samples below the value required for drinking water under French law. Fig. 1 shows the CLD concentration normalized by the initial concentration at different processing times, plotted in log units. The left subfigure (Fig. 1A) shows the surface water (Exp #1 and 2), and the right subfigure (Fig. 1B) shows the deionized water spiked at a similar concentration in CLD. A linear trend line is fitted to the experimental data, with its origin forced to zero, and the slope gives the removal rate constant according to the first-order kinetic model. The aqueous matrix has no very significant effect on the removal rate of CLD. The surface water samples exhibit a removal rate constant of -0.618 and -0.708 $h^{-1}$. On the other hand, we obtain -0.594 and -0.612 $h^{-1}$ for the spiked deionized water. Exp. #1, which corresponds to the smallest initial concentration (0.30 µg L$^{-1}$), also corresponds to the slowest removal ($k$ = -0.618 $h^{-1}$) for the surface water sample. The initial concentration of CLD in the deionized water is very similar between the two samples, as is the degradation rate constant. Overall, the surface water sonolysis shows slightly faster removal compared to deionized water. Notably, the first-order kinetic model does not perfectly fit the experimental data for DI water. Although the coefficient of determination is high ($R^2$ ≈ 0.97), the removal lags behind the model predictions at early time points and accelerates towards the end of the processing time. This occurred in both experiments with deionized water at a lower concentration. In other words, the removal is slow at the beginning, but the overall kinetics are close because the rate increases later. This behaviour could be due to the presence of organic or inorganic scavengers. As demonstrated in several recent studies on the sonolysis of per- and polyfluoroalkylated substances, some salts, like bicarbonate or sulfate, can inhibit the removal, and some organic molecules, such as volatile organic compounds, can compete for the reactive sites, hence decreasing the removal yield and rate (Zeidabadi et al., 2023). However, this is not the case here, as it concerns deionised rather than natural water samples. Another reason could be due to the sonolysis itself, i.e., if the degradation is mediated by reactive species such as $\cdot OH$, which takes time to generate in sufficient quantities, especially if the pollutant is present at a low concentration. However, this phenomenon should also be observed in the surface water sample. CLD is a hydrophobic molecule, so it is more likely that the primary degradation mechanism involves adsorption to the cavitation bubbles, followed by thermal decomposition. At low concentrations, the pollutant has a lower probability of encountering bubbles, especially in deionized water, where no colloidal particles can catalyse bubble formation (nucleation sites). Indeed, cavitation can be described in three stages, nucleation, bubble growth, and collapse. However, it is known that nucleation cannot arise spontaneously from the expansion of the acoustic wave in pure liquid, but rather from particulate matter or microbubbles from previous cavitation events (Suslick et al., 1990). This reasoning can explain why the CLD removal from DI water is slow at the beginning (lack of nucleation sites) and is consistent with the known mechanisms underlying the sonochemical degradation of hydrophobic compounds. Once the cavities are created, previous cavities can be used as nucleation sites, hence the removal rate accelerating. In terms of application, it highlights that the colloidal organics or inorganics constituents in our natural water samples are not scavengers but rather likely enhance sonolysis by catalyzing bubble formation and growth, thereby accelerating the removal of CLD. Total suspended solids measured 5.9 mg/L in the surface water samples, which is not particularly high for surface water. To be more specific, we conducted calorimetry analyses to distinguish whether the likelihood of contact is increased because more cavitation bubbles are created or because CLD is preferentially bound to particulate matter, which serves as nucleation sites. If more cavitation bubbles are created, the temperature elevation in time should be higher, which turns out not to be the case. The DI water shows a temperature increase at a rate of 1.13 $^\circ C \min^{-1}$, whereas the surface water samples reached a rate of 1.03 $^\circ C \min^{-1}$. This suggests that the slightly faster rate of CLD removal in surface water is likely due to

<figure><img src="image_4.png" /><figcaption>**Fig. 1.** Removal of CLD from surface water (left) and deionized spiked water (right). The removal is expressed as the concentration at a given time normalized by the initial concentration. The ratio is expressed in log units, following the first-order kinetic model. Markers show the experimental data, and lines show the fitted linear trend. The degradation is similar between both matrices.</figcaption></figure>
the preferential binding of CLD to suspended solids, which then serve as cavitation bubble nucleation sites, thereby optimizing the contact between the hot spots and the pollutant, resembling the already observed and provoked “heterogeneous” cavitation (Jonnalagadda et al., 2021; Mahendran et al., 2025). This analysis does not strictly rule out the importance of free radicals in the CLD degradation, which could be increased by the presence of metals in the surface water (Yim et al., 2003); or the reactive chlorine species generated during CLD dechlorination, as exhibited for sonolysis of carbon tetrachloride (Dehane et al., 2021).

### 4.3. CLD concentration effect

According to the experiments, an initial CLD concentration about ten times higher increases the removal rate constant by 22 %. This is illustrated in Fig. 2, where the CLD concentration normalized by the initial concentration is plotted against time in log units. The left subfigure (Fig. 2A) displays deionized water samples with an initial concentration of approximately 4 µg L⁻¹ (Exp #3 and #5), while the right subfigure (Fig. 2B) shows samples with an initial concentration of around 0.4 µg L⁻¹ (Exp #4 and #6). A linear trend line is fitted to the experimental data with the origin forced to zero; its slope represents the removal rate constant based on the first-order kinetic model. The experimental results for the higher initial CLD concentration closely follow a first-order kinetic model, with removal rate constants of −0.720 and −0.750 h⁻¹. This indicates that even if the concentration is higher, it is not high enough to saturate the reactive sites (i.e., bubbles). It also reinforces the earlier interpretation that the limiting factor at low concentrations is the probability of CLD encountering cavitation bubbles. Indeed, at higher CLD concentration, as shown in the left subfigure, the first-order kinetics is well followed without any lag as observed at early time on the lower concentration – left subfigure. Overall, the results suggest that thermal decomposition within the cavitation bubble is the dominant factor contributing to CLD removal by sonolysis.

### 4.4. Energy and upscaling considerations

The results of all experiments are summed up in Table 2. Based on the removal rate constant, it is possible to calculate the EEO, following Eq. 2. The EEO is given in Table 2. The values range from 444 to 561 kW h m⁻³, with the lowest value being obtained for Exp #5 (DI water, C₀ = 3.45 µg L⁻¹), and the highest value for Exp #4 (DI water, C₀ = 0.41 µg L⁻¹). However, even the highest EEO is three times lower than the EEO reported for CLD photolysis (Cruz-González et al., 2018). It is of interest that the treatment of surface water contaminated at a “low” concentration, such as 0.3 µg L⁻¹, can still be treated to remove CLD and bring it below the drinking water standard under French law (0.1 µg L⁻¹) at operational costs that are relatively low for this technology. Indeed, sonolysis is often considered an energy-intensive technique. Consequently, ultrasound-assisted removal of CLD is particularly interesting for treating concentrated liquids (e.g., effluents from solid-liquid extraction methods), allowing for the full benefit of the technique. However, it should be noted that although the EEO decreases with increasing CLD concentration, the relation cannot hold for very large concentrations, as all reactive sites may be saturated at some point, thereby decreasing the removal rate. Consequently, given that the upscaling of reactor size is notoriously challenging, the path to pilot-scale application is to identify the maximum CLD concentration above which the EEO decreases and then to integrate it into a treatment chain that includes an upstream concentration step. The addition of suspended solids prone to bind with CLD for promoting heterogeneous cavitation should be considered.

<figure><img src="image_5.png" /><figcaption>**Fig. 2.** Removal of CLD from deionized spiked water at high initial concentration (A - left) and deionized spiked water at low initial concentration (B - right). The removal is expressed as the concentration at a given time normalized by the initial concentration. The ratio is expressed in log units, following the first-order kinetic model. Markers show the experimental data, and lines show the fitted linear trend. The degradation is slightly faster when the CLD concentration is higher.</figcaption></figure>
<table><caption>**Table 2**
Degradation rate constant, R² of the linear regression, and EEO for all experiments.</caption><thead><tr><th>Exp #</th><th>Matrix</th><th>Initial CLD concentration (µg L<sup>-1</sup>)</th><th>Removal after 2 h (%)</th><th>k (h<sup>-1</sup>)</th><th>R<sup>2</sup></th><th>EEO (kW h m<sup>-3</sup>)</th></tr></thead><tbody><tr><td>1</td><td>Surface water</td><td>0.30</td><td>73</td><td>-0.618</td><td>0.97</td><td>539</td></tr><tr><td>2</td><td>Surface water</td><td>0.36</td><td>75</td><td>-0.708</td><td>0.99</td><td>471</td></tr><tr><td>3</td><td>DI water</td><td>3.99</td><td>76</td><td>-0.720</td><td>0.99</td><td>463</td></tr><tr><td>4</td><td>DI water</td><td>0.41</td><td>72</td><td>-0.594</td><td>0.96</td><td>561</td></tr><tr><td>5</td><td>DI water</td><td>3.45</td><td>78</td><td>-0.750</td><td>0.99</td><td>444</td></tr><tr><td>6</td><td>DI water</td><td>0.41</td><td>72</td><td>-0.612</td><td>0.97</td><td>544</td></tr></tbody></table>

## 5. Conclusion

By comparing CLD removal from natural surface water and deionized contaminated water at relevant CLD concentrations, this study demonstrates that sonolysis can degrade 72–78 % of CLD within 2 h, regardless of the water matrix. Rather than acting as scavengers, organic and/or inorganic constituents present in natural surface water (in this case, from the Pérou River in the French West Indies) may actually enhance sonolysis. They likely do so by catalyzing bubble formation and growth, thereby increasing the chances of the contaminant encountering cavitation bubbles. This likelihood is also increased at higher CLD concentrations, with a removal rate constant 22 % higher when the concentration is increased from 0.4 µg L⁻¹ to 4 µg L⁻¹. The tests were conducted on a batch trial of 300 mL, which is far from a sustainable industrial application. However, the energy consumption of the method proved to be more favorable than that of photolysis or photo-Fenton degradation by a factor of 3, although the CLD concentrations differed, suggesting that upscaling efforts are worthwhile. The potential degradation products were not analyzed, and since the complete mineralization of other organochlorine pesticides by sonolysis has not been systematically reported, an important continuation of this work is to study potential by-product formation, their rate of removal, and their potential role in the CLD removal itself.

### CRediT authorship contribution statement

**Stéfan Colombano:** Writing – review & editing, Supervision, Methodology, Conceptualization.
**Deborah E. Martin:** Writing – original draft, Methodology, Investigation.
**Maxime Cochennec:** Writing – original draft, Methodology, Investigation, Conceptualization.
**Alexandre Baudouard:** Resources.
**Benjamin Laulier:** Writing – review & editing, Resources.
**Eric D. van Hullebusch:** Writing – review & editing, Supervision, Methodology.
**Pascal Tierce:** Resources.
**Julien Parinet:** Writing – original draft, Supervision, Methodology, 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.

### Acknowledgments

The authors thank the OPALE Observatoire, and specifically Dr. Antoine Richard from INRAE, for providing them with the contaminated water. The authors also thank Mr. Charles Cousin for the natural water samples and Mr. Thibauld Conte for feedback on the analytical aspect. The authors affiliated with BRGM would like to thank the PIVOTS programme funded by the Région Centre-Val de Loire and the PRIME platform.

### Data availability

Data will be made available on request.

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