# Investigation of PFAS/nPd/µZVI interactions: mechanism, performance, and nPd/µZVI reuse

Raphaël Tur, Stéphanie Betelu, Stéfan Colombano, Dorian Davarzani, Sébastien Bristeau, Julien Grandclément, Arnaud Perrault, Julie Lions, Eric D. van Hullebusch, Ioannis Ignatiadis

► To cite this version:

Raphaël Tur, Stéphanie Betelu, Stéfan Colombano, Dorian Davarzani, Sébastien Bristeau, et al. Investigation of PFAS/nPd/µZVI interactions: mechanism, performance, and nPd/µZVI reuse. Science of the Total Environment, 2025, 1001, pp.180445. 10.1016/j.scitotenv.2025.180445. hal-05285075

HAL Id: hal-05285075

https://brgm.hal.science/hal-05285075v1

Submitted on 26 Sep 2025

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# Investigation of PFAS/nPd/µZVI interactions: mechanism, performance, and nPd/µZVI reuse

Raphaël Tur<sup>a,b,c,*</sup>, Stéphanie Betelu<sup>a</sup>, Romain Rodrigues<sup>a</sup>, Stéfan Colombano<sup>a</sup>, Dorian Davarzani<sup>a</sup>, Sébastien Bristeau<sup>a</sup>, Julien Grandclément<sup>b</sup>, Arnaud Perrault<sup>b</sup>, Julie Lions<sup>a</sup>, Eric D. van Hullebusch<sup>c</sup>, Ioannis Ignatiadis<sup>a,**</sup>

## HIGHLIGHTS

* Sorption dominates PFOA/PFOS removal by nPd/µZVI at pH 4.6–14.

* Anoxic galvanic corrosion of nPd/µZVI controls PFAS sorption on particles.

* PFAS sorption is driven by surface complexation and hydrophobic interactions.

* nPd/µZVI can be used for 4–5 cycles without significant loss in removal efficiency.

## GRAPHICAL ABSTRACT

<figure><img src="image_4.png" /></figure>

## ARTICLE INFO

**Keywords:**

nPd/µZVI particles

PFAS abatement

Corrosion

Sorption

Electrostatic interactions

Surface complexation

Hydrophobic interactions

## ABSTRACT

Due to their persistence, toxicity, and environmental spread, PFAS have been extensively studied for removal since the early 2000s. This study investigates the interactions between nano-sized palladium-coated zero-valent iron microparticles (nPd/µZVI) and perfluorooctanoic acid (PFOA) or perfluorooctanesulfonic acid (PFOS) to examine the mechanism and evaluate the performance of these two PFAS abatement.

Batch experiments were performed under anoxic conditions (1–72 h, 90 °C) using 1 g of ultrasonicated (45 Hz) slurries of nPd/µZVI and 9.29 to 485 mg L⁻¹ of PFAS across a pH range of 1–14. After the reaction and subsequent settling, samples of the aqueous phase were collected for pH measurement, fluoride analysis using a fluoride ion-selective electrode (F-ISE), and quantification of residual PFOA or PFOS and potential byproducts using ultra-high-performance liquid chromatography coupled with mass spectrometry (UHPLC-MS). Methanol used to wash nPd/µZVI under oxic conditions was also analyzed to differentiate between PFAS adsorption and their defluorination. The washing step enabled reuse of the nPd/µZVI, which were characterized by XRD and SEM-EDS before interaction, after interaction, and following methanol washing.

<footnote><sup>a</sup> BRGM, 3 avenue Claude-Guillemin - BP 36009, 45060, Orléans, France</footnote>

<footnote><sup>b</sup> Colas Environnement, 91, rue de la Folliouse, 01700, Miribel, France</footnote>

<footnote><sup>c</sup> Université Paris Cité, Institut de physique du globe de Paris, CNRS, F-75005, Paris, France</footnote>

<footnote>* Corresponding author.</footnote>

<footnote>** Correspondence to: I. Ignatiadis, BRGM, 3 avenue Claude Guillemin, BP 36009, 45060 Orleans, France
E-mail addresses: raphael.tur@cer.colas.fr (R. Tur), i.ignatiadis@brgm.fr (I. Ignatiadis).</footnote>
PFAS sorption, governed by surface complexation and hydrophobic interactions, is the predominant PFAS abatement mechanism. It is driven by the anoxic galvanic corrosion of nPd/µZVI (AGC-nPd/µZVI), which modifies the physical and chemical conditions of the liquid phase and the surface chemistry of nPd/µZVI as a function of pH. Under the investigated conditions, a linear sorption model accurately described PFAS abatement at pH 4, where nPd/µZVI were shown to be reusable for at least 4–5 cycles without significant loss of sorption efficiency.

# 1. Introduction

Per- and polyfluoroalkyl substances (PFAS) are a large class of synthetic chemicals, comprising between 5000 and over 40,000 distinct compounds. PFAS carry at least one fully fluorinated carbon and one functional group. Carboxylates (PFCAs, C₆F₂nCOOH with n < 7 for short-chain and n ≥ 7 for long chain), sulfonates (PFSAs, C₆F₂nSO₃H with n < 6 for short-chain and n ≥ 6 for long chain) acids, or perfluoroalkyl acids (PFAA) precursors such as fluorotelomer-based substances C₆F(2n+1)-C₂H₄-R (Environ. Sci. Technol. Lett., n.d.), have been widely used in industrial and consumer applications since the 1950s (Glüge et al., 2020). PFAS have been primarily used in applications that require extremely low surface energy or surface tension, as well as durable water- and oil-repellent properties, such as in fire-fighting foams, surface treatments for textiles, carpets, and papers, as well as in shampoos and paints. Today, PFAS are ubiquitous in the environment (Abunada et al., 2020). Worldwide PFAS contamination in soil, groundwater, and surface water varies considerably, with environmental levels ranging from 10 to 100 ng kg⁻¹ to over 10 µg kg⁻¹, and rarely exceeding 1 mg L⁻¹ in natural waters (e.g., 1–94 ng L⁻¹ for PFAA (EWG, n.d.)) or in tap waters (approximately 80 ng L⁻¹) (Ingold et al., 2023). A European spatial study found PFAS concentrations in soil from 17 ng kg⁻¹ to 5350 mg kg⁻¹, identifying 1938 high-level sites (≥10 µg kg⁻¹) grouped into 288 clusters (Moghadasi et al., 2023; Jiang et al., 2025). Their strong carbon-fluorine (C—F) bonds, characterized by high electronegativity and chemical stability, greatly limit their natural attenuation (Newell et al., 2021). Many studies have demonstrated their adverse health effects (Abunada et al., 2020). In 2023, the IARC classified perfluorooctanoic acid (PFOA) and perfluorooctanesulfonic acid (PFOS) (Fig. 1), two of the most widely used PFAS, as carcinogenic (group 1) and possibly carcinogenic (group 2B) to humans, respectively (Zahm et al., 2024).

In China, PFOS levels reached 778 mg kg⁻¹ in soil and 1.74 mg L⁻¹ in groundwater (Jiang et al., 2025). In the U.S., aqueous film-forming foam (AFFF)-contaminated soils showed PFOS levels up to 460 mg kg⁻¹ and PFOA over 50 mg kg⁻¹, with a median PFOS concentration of 8.72 mg kg⁻¹ (Brusseau et al., 2020). Industrial products like firefighting foams can contain PFAS concentrations ≥100 mg L⁻¹, with undiluted AFFF

reaching 4900–11,400 mg L⁻¹ (Backe et al., 2013). While limiting their use, it is thus crucial to design treatment processes capable of operating effectively over a wide concentration range spanning several orders of magnitude.

Among the treatment technologies investigated (Usman et al., 2024; Askeland et al., 2020; Sörengård et al., 2020; Cheng et al., 2024; Dixit et al., 2021; Vargette et al., 2023; Trautmann et al., 2015; Xia et al., 2022; Rodriguez-Freire et al., 2015; Krause et al., 2022; Hao et al., 2021; Trang et al., 2022), zero-valent iron particles (ZVI) have attracted interest for PFAS remediation (Blotevogel et al., 2018; Zhang et al., 2018; Fang et al., 2024a) due to their effectiveness in the reductive degradation of chlorinated organic solvents (Rodrigues et al., 2020). However, this effectiveness is contingent upon minimizing, in the presence of O₂, the precipitation of Fe(III) oxyhydroxides onto the ZVI, which can decrease the reductive capacity of ZVI, e.g., by using a combination of ZVI and sodium dithionite (Rodrigues et al., 2020). In terms of redox processes between ZVI and PFAS, an experimental–theoretical study demonstrated that the reductive dehalogenation of PFOA is strongly limited by kinetic constraints, unless suitable catalysts are employed (Blotevogel et al., 2018). In fact, with nanoscale nZVI, defluorination of PFOA and PFOS was found to be ineffective; however, significant removal was achieved via sorption, primarily driven by hydrophobic and electrostatic interactions (Zhang et al., 2018; Fang et al., 2024a).

Sorption on ZVI depends on the nature of interfacial iron corrosion products, Fe(II) and Fe(III) species that dissolve, hydrolyze, precipitate and or re-dissolve due to amphoteric behavior, depending on the chemical and physical conditions, including redox conditions, temperature, and pH. For example, under oxic conditions, the sorption capacity of zero-valent iron particles (ZVIs) gradually decreases as iron corrosion products precipitate in the form of Fe(III) oxyhydroxide (Zhang et al., 2018). Indeed, ferrihydrite (Fe(OH)₃) starts to precipitate at pH values above 1.8 at 25 °C, and evolves into goethite (FeOOH) according to reactions described in Eqs. (1) and (2) (Schwertmann and Fechter, 1982; Cuss et al., 2020).

$$3\text{O}_2 + 6\text{H}_2\text{O} + 4 \text{Fe} \rightarrow 4 \text{Fe(OH)}_3 \quad (1)$$

<figure><img src="image_5.png" /><figcaption>Fig. 1. Three-dimensional representations of PFOA and PFOS (generated using ChemAxon software, Budapest, Hungary).</figcaption></figure>
$$3O_2 + 2H_2O + 4 Fe \rightarrow 4 FeOOH + 3H^+ \quad (2)$$

FeOOH is the predominant form in the pH range 1.8–14 at 25 °C, in equilibrium with dissolved species, Fe³⁺ until pH 2.5, FeOH²⁺ until pH 3.5, Fe(OH)₂ until pH 6.8, Fe(OH)₃ until pH 9, and Fe(OH)₄ until pH 14.0. The point of zero net charge (pHₚZNC) of synthetic Fe-oxyhydroxides ranges between pH 6.8 and 9 (Duquesne et al., 2020a; Duquesne et al., 2020b; Duquesne et al., 2022). Thus, the nature and the predominance of interfacial species control the surface charge of ZVI, according to Brønsted–Lowry acid-base theory.

PFAS sorption efficiency on ZVI depends on the carbon-fluor chain length and functional group. It decreases with shorter CF₂ chains and increased for sulfonate head groups compared to carboxylate groups of the same chain length (Zhang et al., 2018; Fang et al., 2024a). Sulfonate head groups showed higher sorption, likely due to stronger hydrophobic and electrostatic interactions, as well as higher adsorption energy (Fang et al., 2024b).

Bimetallic Pd/ZVI particles have been investigated for PFAS defluorination (Lawal and Choi, 2018; Park et al., 2018) owing to palladium's catalytic role in promoting dehalogenation, under acidic and anoxic conditions (Rodrigues et al., 2019). While pure Pd nanoparticles combined with H₂ bubbling enabled PFOA defluorination, releasing fluoride ions accounting for 46 % of the total fluorine from PFOA within 36 h at 25 °C (Long et al., 2021), bimetallic Pd/ZVI particles primarily removed PFOA and PFOS through sorption, as no significant fluorinated by-products were detected, indicating limited reductive defluorination (Lawal and Choi, 2018; Park et al., 2018). However, Pd/ZVI exhibited higher removal efficiency than ZVI (Lawal and Choi, 2018), likely due to its higher pHₚZNC, enhancing electrostatic interactions, and the potential contribution of palladium to nonreactive PFOA chemisorption (Long et al., 2021).

Building on our previous results on recalcitrant chlorinated compounds (Rodrigues et al., 2019) and those of bibliography on fluorinated compounds (Lawal and Choi, 2018; Park et al., 2018; Long et al., 2021), in this study, we investigated, under anoxic conditions, the interactions between nano-sized palladium-coated zero-valent iron microparticles (nPd/µZVI) and PFOA or PFOS with nPd/µZVI vs PFAS molar ratios ranging from 1450 to 91,600, at 90 °C, to maximize the galvanic corrosion of the iron particles and enhance H₂ production to promote defluorination. The PFAS concentrations ranging from 9.29 to 455 mg L⁻¹ were selected to analyze the remaining PFOA or PFOS and potential by-products at the µg L⁻¹ level, using UHPLC-MS, while keeping optimum conditions for measuring fluoride ions at the mg L⁻¹ level, using a fluoride-ion selective electrode (F-ISE). The interactions between nPd/µZVI and PFOA or PFOS were investigated to assess the mechanism and performance of PFAS abatement, while the influence of C–F chain length was further examined using perfluorohexanesulfonic acid (PFHxS) and perfluorobutane sulfonic acid (PFBS). Batch experiments were conducted under anoxic conditions (1–72 h, 90 °C) using 1 g of ultrasonicated (45 Hz) nPd/µZVI slurries across a pH range of 1–14. After the reaction and subsequent settling, samples of the aqueous phase were collected for pH measurement and for the analysis of fluoride, residual PFOA or PFOS, and potential by-products concentrations. Methanol used to wash nPd/µZVI under oxic conditions was also analyzed to differentiate between PFAS adsorption and their defluorination. This washing step further enabled the reuse of nPd/µZVI, which were characterized by XRD and SEM-EDS before interaction, after interaction, and following methanol washing.

## 2. Materials and methods

### 2.1. Chemicals

Perfluorooctanesulfonic acid potassium salt C₈F₁₇SO₃K (linear K-PFOS, 99.8 wt%) was supplied by TechLab (Saint-Julien-Lès-Metz, France), perfluorohexanesulfonic acid C₆F₁₃SO₃H (PFHxS, technical

grade) from LGC (Teddington, Middlesex, UK) and perfluorobutanesulfonic acid C₄F₉SO₃H (PFBS) from Apollo Scientific (Manchester, UK). Perfluorooctanoic acid C₇F₁₅COOH (PFOA, 95 wt%), anhydrous dimethyl sulfoxide C₂H₆OS (anhydrous DMSO, > 99.9 %), and sodium hydrogenocarbonate (NaHCO₃, 99.5–100.5 %) were purchased from Sigma-Aldrich (Saint Quentin Fallavier, France). Hydrochloric acid (HCl, 35–38 wt%), acetic acid (CH₃COOH, > 99 %), and sodium carbonate (Na₂CO₃, 100 %) were supplied by Fisher Chemical (Illkirch, France). Sodium hydroxide (NaOH, 99.1 wt%) was provided by VWR Chemicals (Rosny-sous-Bois, France). Methanol MeOH (Optima LC/MS grade) was provided by Fisher Chemical.

For analysis, PFAS standards and internal standards were purchased from Wellington (Ontario, Canada) and Neochema (Darmstadt, Germany).

Ultrapure water (HPLC grade) and methanol (optima LC-MS grade) were purchased from Fisher Chemical. Ammonium acetate (purity 99 %) and glacial acetic acid (purity 99 %) were purchased from Merck (Darmstadt, Germany), ammonia solution (25 %, analytical reagent grade) was purchased from Fisher Chemical.

Nano-sized palladium-coated zero-valent iron microparticles (nPd/µZVI) were provided by Bioren Engineering (Ghent, Belgium). According to the manufacturer and previous studies, the Pd content is less than 5 wt% (Rodrigues et al., 2019; Dien et al., 2013).

### 2.2. Solutions

All aqueous solutions were prepared in Milli-Q water (18.2 MΩ cm at 25 °C) provided by a Milli Q Integral 3 system (Merck Millipore, Guyancourt, France). K-PFOS, PFHxS, PFBS and PFOA stock solutions (5 g L⁻¹) were prepared by dissolving the required mass of solid in DMSO. Experiments were conducted in 0.1 M HCl (pH 1), acetate buffer solutions (pH 4 and 5; ionic strength (IS) = 0.1 M), carbonate buffer solutions (pH 9, 10 and 11, IS = 0.1 M), 1 M NaOH (pH 14) and in a commercial drinking water (Eau de Source, Saint Hippolyte, France): pH 7.8, Ca²⁺ = 63 mg L⁻¹; HCO₃⁻ = 164 mg L⁻¹; Mg²⁺ = 11 mg L⁻¹; K⁺ = 4 mg L⁻¹; Na⁺ = 14 mg L⁻¹; NO₃⁻ < 3 mg L⁻¹. Commercial drinking water was used to obtain a near-neutral pH solution containing carbonate anions, like buffer solutions, without resorting to phosphate-based or organic buffers.

### 2.3. Batch experiments

Glass vials of 20 mL with aluminum Cap, equipped with a Silicone/ PTFE septum (20 mm Crimp Cap Vial Kit, Headspace, Clear Glass) were purchased from Fisherbrand (Illkirch, France). The ultrasonic tank (Transsonic TI-HS-5, 25–45 kHz, 550 W, 3.5 L) was provided by Elma Schmidbauer GmbH (Singen, Germany). Experiments were conducted across a pH range of 1 to 14, in 20 mL sealed batches using 10 mL of slurries containing 1 g of nPd/µZVI and 9.29 to 485 mg L⁻¹ of PFAS, ultrasonically homogenized at 90 °C, in the range from 3 to 72 h.

The amount of iron was sufficiently high to reduce all initially dissolved and sealed gaseous oxygen under an air atmosphere, thereby forcing anoxic corrosion, with a molar ratio of total iron to oxygen higher than 214: 214 ≤ Feₜot/(O₂dissolved + O₂gas)ₜot ≤ 220, considering that 2–5 % of the solid material consists of Pd and secondary phases ((O₂dissolved + O₂gas)ₜot = 7.97 × 10⁻⁵ mol; 1.70 × 10⁻² ≤ Fe (mol) ≤ 1.75 × 10⁻²).

General and basic knowledge on the corrosion of Fe⁺ under anoxic conditions are presented hereafter (Schwertmann and Fechter, 1982; Cuss et al., 2020). Under anoxic conditions, Fe corrodes, according to (Eqs. (3a), (3b), (4) and (5)):

$$2H^+ + Fe \rightarrow Fe^{2+} + H_2 \quad (3a)$$

$$2H_2O + Fe \rightarrow Fe^{2+} + 2OH^- + H_2 \quad (3b)$$
Dissolved Fe²⁺ is the predominant form in the pH range 0 to 8.8 at 25°C.

$$2\text{H}_2\text{O} + \text{Fe} \rightarrow \text{Fe(OH)}_2 + \text{H}_2 \quad (4)$$

For pH higher than 8.8, Fe(OH)₂ precipitates. Fe(OH)₂ is the predominant form in the pH range until 13.8, in equilibrium with dissolved species, Fe²⁺ until pH 8.5, FeOH⁺ until pH 11.2, Fe(OH)₂ until pH 11.3, Fe(OH)₃ until pH 13.8. Above pH 13.8, the predominant Fe(II)-form is FeO₂⁻, according to Eq. (5):

$$2\text{OH}^{-} + \text{Fe} \rightarrow \text{FeO}_{2}^{-} + \text{H}_{2} \quad (5)$$

Table 1 presents the nature of the experiment as well as the associated conditions conducted across a pH range of 1 to 14. Each experiment was performed in a new glass vial and conducted in triplicate. After interaction, vials were allowed to settle for 24 h at room temperature to separate nPd/µZVI from the liquid phase. nPd/µZVI particles visibly settled within 24 h when the interaction experiments were conducted using an aqueous phase with pH 1–11. At pH 14, residual suspended nPd/µZVI caused turbidity in the samples after PFAS-nPd/µZVI interaction, making further sampling more difficult. The observed contact angle increased at the solid-water-air interface, meaning that nPd/µZVI surface was hydrophobic, i.e., PFAS oriented its polar groups toward the surface of nPd/µZVI and its fluorocarbon chains outward. Glass vials were not adapted for the centrifugation process at 10,000g. 9 mL of the supernatant were then carefully sampled, transferred into a 15 mL polypropylene tube and centrifuged at 10,000g (Sigma 3–30 K, Osterode am Harz, Germany) for 10 min before sampling 1 mL of the liquid phase for chemical analysis.

Although thermodynamic equilibrium was eventually reached, the 24-hour period was not considered the reaction time due to slowed kinetics caused by the reduced temperature. Nevertheless, the results still reflect equilibrium at 25 °C. The remaining solid in the vials underwent

<table><caption>**Table 1**

Matrix of experiments conducted for an in-depth investigation of PFAS – nPd/µZVI interactions. C₀ is the PFAS initial concentration.</caption><thead><tr><th>N°</th><th>Nature and objective of experiment</th><th>Experimental conditions</th></tr></thead><tbody><tr><td>1</td><td>Data accuracy</td><td><ul><li>9.29 ≤ C₀ (mg L⁻¹) ≤ 485</li><li>V = 10 mL; pH = 4</li><li>T = 90 °C (in ultrasonic bath)</li><li>Reaction time: 72 h</li></ul></td></tr><tr><td>2</td><td>Investigation of the abatement mechanism of PFAS using nPd/µZVI</td><td><ul><li>C₀(PFOS) = 10 mg L⁻¹; C₀(PFOS) = 9.29 mg L⁻¹</li><li>nPd/µZVI = 1 g (100 g L⁻¹)</li><li>V = 10 mL; 1 ≤ pH ≤ 14</li><li>T = 90 °C (in ultrasonic bath)</li><li>Reaction time: 72 h</li></ul></td></tr><tr><td>3</td><td>PFAS removal kinetics</td><td><ul><li>C₀(PFOS) = 10 mg L⁻¹; C₀(PFOS) = 9.29 mg L⁻¹</li><li>nPd/µZVI = 1 g (100 g L⁻¹)</li><li>V = 10 mL; pH = 4</li><li>T = 90 °C (in ultrasonic bath)</li><li>Reaction time: 1 to 72 h</li></ul></td></tr><tr><td>4</td><td>Influence of PFAS-nPd/µZVI ratio</td><td><ul><li>9.29 ≤ C₀ (mg L⁻¹) ≤ 485</li><li>nPd/µZVI = 1 g (100 g L⁻¹)</li><li>V = 10 mL; pH = 4</li><li>T = 90 °C (in ultrasonic bath)</li><li>Reaction time: 3 h</li></ul></td></tr><tr><td>5</td><td>Influence of carbon-fluor chain length</td><td><ul><li>C₀(PFOS) = 9.29 mg L⁻¹; C₀(PFOS/ PFHXS) = 10 mg L⁻¹</li><li>nPd/µZVI = 1 g (100 g L⁻¹)</li><li>V = 10 mL; pH = 4</li><li>T = 90 °C (in ultrasonic bath)</li><li>Reaction time: 3 h</li></ul></td></tr><tr><td>6</td><td>nPd/µZVI reuse</td><td><ul><li>C₀(PFOS) = 10 mg L⁻¹; C₀(PFOS) = 9.29 mg L⁻¹</li><li>nPd/µZVI = 1 g (100 g L⁻¹)</li><li>T = 90 °C (in ultrasonic bath)</li><li>V = 10 mL; pH = 4</li><li>Reaction time: 3 h</li></ul></td></tr></tbody></table>

a methanol extraction process to quantify PFAS sorbed fraction onto nPd/µZVI (Park et al., 2018). It involved: (i) adding 10 mL of methanol, (ii) vortexing the slurry for a few seconds, (iii) interacting nPd/µZVI vs PFAS, (iv) decanting particles for 24 h, and (v) carefully sampling 10 mL of the supernatant. The procedure was repeated up to five times to recover up to 50 mL of sample, which was collected into a 50 mL polypropylene tube before sampling 1 mL for chemical analysis. Prior to reuse, the nPd/µZVI materials were washed three times with methanol and subsequently dried for 72 h under ambient laboratory conditions. One cycle consists of performing the PFAS/nPd/µZVI interaction, followed by nPd/µZVI washing and drying procedures. Before drying, 1 mL of methanol from washing remained in the vial. During one PFAS/nPd/µZVI interaction, the complete redox reaction between O₂ (O₂dissolved + O₂gas)tot and nPd/µZVI leads to the production of 1.06 × 10⁻⁴ mol of Fe (III). One PFAS/nPd/µZVI interaction consumes 0.61–0.62 % of the initial iron amount for O₂ reduction, two cycles consume 1.21–1.25 %, and eight cycles consume 4.84–4.99 %, considering that 95–98 % of the solid material consists of µZVI.

## 2.4. Analysis

### 2.4.1. Liquid phases

#### 2.4.1.1. UHPLC-MS
PFAS analyses used a Waters TQXS system (Guyancourt, France) in conjunction with a Waters UHPLC system, featuring an Acuity BEH C18 Column (100 × 2.1 mm, 1.7 µm, Waters) maintained at 35°C, along with a delay column (Atlantis Premier BEH C18 AX 50 mm × 2.1 mm, 5 µm, Waters) to prevent PFAS contamination from the mobile phase and the chromatographic system. The mobile phase was a mixture of 2 mM ammonium acetate in water (solution A) and 2 mM ammonium acetate in methanol (solution B). Table 2 shows the temporal evolution of the solvent composition (elution gradient).

The TQXS mass spectrometer was utilized for MS analysis, functioning in negative Electrospray Ionization (ESI) mode. The source conditions were established as follows: desolvation temperature of 500°C, desolvation gas flow at 1100 L hr⁻¹, cone gas flow at 150 L hr⁻¹, and capillary voltage set to –1000 V (ESI–). The chromatograms underwent processing using Target Lynx software. The mass spectral parameters used for PFAS analysis are detailed in Table 3.

The different PFAS quantified were perfluorobutanoic acid (PFBA, C₃F₇COOH), perfluoropentanoic acid (PFPeA, C₄F₉COOH), perfluorohexanoic acid (PFHxA, C₅F₁₁COOH), perfluoroheptanoic acid (PFHpA, C₆F₁₃COOH), 7H-perfluoroheptanoic acid (HPFHpA, C₆HF₁₂COOH), PFOA (C₇F₁₅COOH), PFBS (C₄F₉SO₃H), PFHxS (C₆F₁₃SO₃H), and PFOS (C₈F₁₇SO₃H). The instrumental limit of quantification is ranging from 2 to 10 ng L⁻¹.

Each 1 mL sampled volume was first diluted by 2353 in a water/methanol mixture (50/50 v/v), in two consecutive steps (by 13 and 181). It was then diluted by 1.14 in a 700 µL polypropylene vial: 440 µL was mixed with 10 µL of 5 % acetic acid and 50 µL of an internal standard solution containing 2 µg L⁻¹ of isotopically labeled PFAS (¹³C₄–PFBA, ¹³C₂–PFHxA, ¹³C₄–PFOA, ¹³C₃–PFBS, ¹⁸O₂–PFHxS and ¹³C₄–PFOS). All these dilutions ensured that PFAS concentrations fell within the calibration range (10 to 5000 ng L⁻¹). The injected volume

<table><caption>**Table 2**

Mobile phase flow rate and composition (solutions A and B) as a function of time during UHPLC analyses.</caption><thead><tr><th>Time (min)</th><th>Flow rate (mL min⁻¹)</th><th>Mobile phase A (% v/v)</th><th>Mobile phase B (% v/v)</th></tr></thead><tbody><tr><td>0</td><td>0.3</td><td>95</td><td>5</td></tr><tr><td>6.5</td><td>0.3</td><td>5</td><td>95</td></tr><tr><td>7.0</td><td>0.3</td><td>5</td><td>95</td></tr><tr><td>7.6</td><td>0.3</td><td>95</td><td>5</td></tr><tr><td>10.0</td><td>0.3</td><td>95</td><td>5</td></tr></tbody></table>
<table><caption>**Table 3**
Mass spectral parameters utilized for PFAS analysis.</caption><thead><tr><th>Acronym</th><th>Ionization mode</th><th>Quantifier m/z (collision energy V)</th><th>Qualifier m/z (collision energy V)</th><th>Cone voltage (V)</th><th>Internal standard</th></tr></thead><tbody><tr><td>PFBA</td><td>ESI –</td><td>213 > 169 (8)</td><td></td><td>10</td><td><sup>13</sup>C4–PFBA</td></tr><tr><td>PFPeA</td><td>ESI –</td><td>263 > 219 (8)</td><td>263 > 69 (35)</td><td>15</td><td><sup>13</sup>C2–PFHxA</td></tr><tr><td>PFHxA</td><td>ESI –</td><td>313 > 269 (6)</td><td>313 > 119 (19)</td><td>15</td><td><sup>13</sup>C2–PFHxA</td></tr><tr><td>PFHpA</td><td>ESI –</td><td>363 > 319 (9)</td><td>363 > 169 (15)</td><td>15</td><td><sup>13</sup>C4–PFOA</td></tr><tr><td>HPFHpA</td><td>ESI –</td><td>345 > 281 (10)</td><td>345 > 131 (23)</td><td>20</td><td><sup>13</sup>C4–PFOA</td></tr><tr><td>PFOA</td><td>ESI –</td><td>413 > 369 (9)</td><td>413 > 169 (17)</td><td>20</td><td><sup>13</sup>C4–PFOA</td></tr><tr><td>PFBS</td><td>ESI –</td><td>299 > 80 (27)</td><td>299 > 99 (26)</td><td>25</td><td><sup>13</sup>C3–PFBS</td></tr><tr><td>PFHxS</td><td>ESI –</td><td>399 > 80 (33)</td><td>399 > 99 (30)</td><td>30</td><td><sup>18</sup>O2–PFHxS</td></tr><tr><td>PFOS</td><td>ESI –</td><td>499 > 80 (37)</td><td>499 > 99 (37)</td><td>30</td><td><sup>13</sup>C4–PFOS</td></tr><tr><td><sup>13</sup>C4–PFBA</td><td>ESI –</td><td>217 > 172 (8)</td><td></td><td>10</td><td></td></tr><tr><td><sup>13</sup>C2–PFHxA</td><td>ESI –</td><td>315 > 270 (9)</td><td>315 > 119 (19)</td><td>15</td><td></td></tr><tr><td><sup>13</sup>C4–PFOA</td><td>ESI –</td><td>417 > 372 (11)</td><td>417 > 169 (17)</td><td>20</td><td></td></tr><tr><td><sup>13</sup>C3–PFBS</td><td>ESI –</td><td>302 > 80 (29)</td><td>302 > 99 (27)</td><td>25</td><td></td></tr><tr><td><sup>18</sup>O2–PFHxS</td><td>ESI –</td><td>403 > 84 (33)</td><td>403 > 103 (30)</td><td>30</td><td></td></tr><tr><td><sup>13</sup>C4–PFOS</td><td>ESI –</td><td>503 > 80 (38)</td><td>503 > 99 (38)</td><td>30</td><td></td></tr></tbody></table>

for analysis was 10 µL.

The pH values of the supernatant were measured using an OGPH207 electrode with an OpH218 meter (OrigaLys, Rillieux-la-Pape, France), calibrated with commercial pH 4, 7, and 10 buffer solutions.

### 2.4.1.2. Fluoride (F⁻) potentiometric analysis using a F⁻ ion-selective electrode (ISE).
The fluoride measurement system consisted of a F-ISE composed of a LaF3 crystal membrane (6.0502.150 model, Metrohm Schweiz AG, Zofingen, Swiss) coupled with a saturated calomel electrode (G0115 SCE model, Ametek (Berwyn, PA, USA) and a titrator model 905 Titrando (Metrohm) interfaced to a PC system with Tiamo 2.0 software (Metrohm). The potentiometric analyses of free fluoride content [F⁻] were performed at room temperature (≈ 23 °C) using NaF standard addition method. The linear domain ranged from 1 to 40 mg L⁻¹. The protocol was adapted to adjust the pH of 5 mL samples to between 5 and 7 using minimal volumes of concentrated HNO₃ or NaOH solutions, in accordance with the operating conditions required for F-ISE. Potentiometric measurements were then conducted by diluting 5 mL of each sample into an appropriate volume of TISAB IV (containing acetic acid; Supelco, purchased from Merck) for fluoride quantification in the 3–35 mg L⁻¹ range. TISAB IV was used to buffer the solution (at pH ≈ 5.6–6), fix the ionic strength, and complex interfering cations to ensure accurate measurement.

#### 2.4.2. Solid phase

Scanning electron microscopy (SEM) observations of nPd/µZVI particles were carried out on a FEG (field emission gun, high resolution) Auriga 40 (resolution: 1 nm, working distance: 8.5 mm, EHT: 15 keV) purchased from Zeiss (Jena, Germany). It was equipped with an Energy-Dispersive X-ray Spectroscopy (EDS) microanalysis system Quanta 200 (resolution: 125 eV) purchased from Bruker (Palaiseau, France). Images were obtained under high vacuum, using a secondary electron detector “In Beam” (IBSE) at 10 kV, and EDS spectra were acquired at 15 kV.

X-ray diffraction (XRD) analysis was made using a D8 Advance da Vinci diffractometer (Bruker) equipped with a copper source (λ = 1.5406 Å) and a 1D LynxEye XE-T detector. The analyses were conducted over an angular range from 4° to 90°2θ, with a step size of 0.03°2θ and a step time of 2.50 s. Mineralogical identification was performed using the DIFFRAC.SUITE software (provided by Bruker) and the ICDD J.C.P.D-S database (2024).

#### 2.5. Calculation of PFAS abatement and sorption isotherm

For each experiment, the percentage of PFAS abatement was calculated by Eq. (6):

$$ \text{PFAS abatement (\%)} = \frac{C_0 - C_{\text{eq}}}{C_0} \times 100 \qquad (6) $$

where C₀ is the PFAS initial concentration (mg L⁻¹) and Cₐₘ is the equilibrium concentration of PFAS in the liquid phase (mg L⁻¹).

The percentage of PFAS desorbed from nPd/µZVI by methanol extraction was calculated by Eq. (7):

$$ \text{PFAS desorbed (\%)} = \frac{C_{\text{methanol}}}{C_0} \times 100 \qquad (7) $$

where Cₐₘ is the PFAS concentration in the methanol extract.

The sorption isotherm was fitted with the linear isotherm (Eq. (8)):

$$ q_e = K_d \times C_{\text{eq}} \times (C_0 - C_{\text{eq}}) \times \frac{V}{m} = K_d \times C_{\text{eq}} \qquad (8) $$

where q_e is the amount of PFAS adsorbed per unit mass of nPd/µZVI at equilibrium (mg g⁻¹), calculated with a liquid/nPd/µZVI ratio (V/m) of 0.01 L g⁻¹, and K_d is the sorption constant (L g⁻¹), representing how much PFAS prefers nPd/µZVI over the liquid phase.

### 2.6. Variability in the dataset

All experiments were performed in triplicate, with one UHPLC-MS analysis conducted per replicate. The variability in UHPLC-MS data ranged from 5 % to 8 % of the measured values. Results are reported as the mean of the triplicates, with error bars representing the standard deviation (n = 3). These error bars do not account for the 5 %–8 % analytical error inherent to UHPLC-MS measurements, nor do they reflect potential dilution errors. It should be noted that from the initial 10 mL of aqueous phase, 9 mL of the supernatant were carefully sampled, leaving 1 mL in contact with the solid phase. This 1 mL residue was analyzed in the total volume of methanol used for washing (see Section 2.3). When sorption predominates on the solid, the amount of PFAS present in this remaining volume is negligible. However, when sorption is limited, the PFAS content in the residual aqueous phase can influence the measurements. In such cases, the percentage of desorbed PFAS may exceed the corresponding removal percentage.

## 3. Results and discussion

### 3.1. Investigation of data accuracy

Fig. 2 presents data accuracy by comparing the theoretical and measured PFAS concentrations at pH 4 and 90 °C. The experiments were conducted with PFAS concentrations ranging from 9.29 to 485 mg L⁻¹ for 72 h in the absence of nPd/µZVI.

The comparison between theoretical and measured values for PFOA and PFOS showed a strong linear correlation (R² = 0.9996 for PFOA and 0.9984 for PFOS) across the 10–485 mg L⁻¹ range. The measurement error was constant, repeatable, and reproducible, 3 % for PFOA and 8 %
for PFOS, indicating minimal influence from the water/air interface ([PFAS] < CMC; CMC<sub>PFOA</sub> = 3214–5716 mg L<sup>-1</sup> and CMC<sub>PFOS</sub> = 536–3501 mg L<sup>-1</sup> (Bhatarai and Gramatica, 2011; Sørli et al., 2020)), no sorption on glass (pH > SiO<sub>2</sub> point-of-zero charge pHPZNC<sub>SiO<sub>2</sub></sub> = 2.0) (Crundwell, 2017), and no degradation under low-frequency sonolysis, in agreement with (Sidnell et al., 2022). Both PFAS remained soluble under the investigated conditions. The consistent error below 10 % confirms data accuracy with regards to the dilution protocol and UHPLC-MS analyses, and all future results will be compared to the 100 % theoretical value.

## 3.2. Investigation of the abatement mechanism of PFAS using nPd/µZVI

### 3.2.1. Experiments conducted at pH 4

Fig. 3 presents the results of the experiments performed at pH<sub>i</sub> = 4 to differentiate between sorption and defluorination mechanisms. nPd/µZVI showed a removal efficiency of 75–85 % for PFOA and ≈ 95–100 % for PFOS. nPd/µZVI successive methanol washings enabled recovering ≈ 90–98 % of the removed PFAS; ≈ 80–85 % after the first washing, ≈ 5–10 % after the second one, and ≤ 1 % from the third. The consistent error below 10 % confirms data accuracy with regards to the removal/washing protocol, successive dilutions, and UHPLC-MS analyses. Under the investigated condition, sorption is the unique removal mechanism of PFOA and PFOS by nPd/µZVI particles. This conclusion is supported by the absence of any detectable organic (see Table 3) and inorganic (F<sup>-</sup>) degradation products, up to the analytical limits of our methods and procedures. For further experiments, only three methanol washings were subsequently carried out to determine the content of sorbed PFAS.

### 3.2.2. Experiments conducted in the pH range 1 to 14

Fig. 4a and Fig. 4b present the influence of pH the removal efficiency (percentage) on PFOA and PFOS by nPd/µZVI particles. According to the error bars, the removal/washing protocol, the successive dilutions, and the UHPLC-MS analyses, sorption is the predominant removal mechanism of PFOA and PFOS by nPd/µZVI particles within the whole investigated pH domain. It is driven by the anoxic galvanic corrosion of nPd/µZVI (AGC-nPd/µZVI), which modifies the physical and chemical conditions of the liquid phase (Fig. 4c) and the surface chemistry of nPd/µZVI as a function of pH (Chivot, 2004). The pH range was further restricted to 4.6 ≤ pH ≤ 14, according to Fig. 4c and Fig. 4d, which illustrate the effective influence of pH on PFOA and PFOS removal by nPd/µZVI particles. For both PFAS, higher sorption efficiency was obtained under acidic conditions. PFAS sorption was highest at pH = 4.6 (~100 %) and remained strong at pH = 5.4 (85–95 % for PFOA, 95–100 % for PFOS). Sorption decreased at pH 6.1 (65 % for PFOA, 80 % for PFOS) and dropped further at neutral to basic pH (11 and 1 % for PFOA

<figure><img src="image_6.png" /><figcaption>**Fig. 2.** Comparison between expected and measured PFAS concentrations in the absence of nPd/µZVI. Conditions: pH = 4, T = 90 °C, 9.29 ≤ [PFAS] (mg L<sup>-1</sup>) ≤ 485.</figcaption></figure>

<figure><img src="image_7.png" /><figcaption>**Fig. 3.** PFAS removal and desorption after 1 to 5 methanol extractions. Conditions: pH = 4, C₀ = 9.29–10 mg L<sup>-1</sup>, nPd/µZVI = 1 g, T = 90 °C, n = 3.</figcaption></figure>

and between 40 and 16 % for PFOS between pH 7.8–11). For both, the minimum adsorption was obtained at pH 9.6. At pH 14, sorption of PFOA and PFOS increased up to 15 % and 60 %, respectively.

In the effective pH range 4 ≤ pH ≤ 10, nPd/µZVI particles visibly settled within 24 h, influenced by their density and hydrophilic surface.

In the pH range 4.0–5.4, the intensity of the anoxic galvanic corrosion of nPd/µZVI (AGC-nPd/µZVI) remains in the same order of magnitude; the massive production of adsorbed Fe<sup>2+</sup> is responsible for the PFAS complexation adsorption process (PFAS-CAP) onto nPd/µZVI. According to Eqs. (3a) and (3b), 2n moles of H<sup>+</sup> are required to produce n moles of Fe<sup>2+</sup>. In the presence of HCl, the amount of H<sup>+</sup> consumed is calculated directly from the initial and final pH values. In the presence of CH<sub>3</sub>COO<sup>-</sup>/CH<sub>3</sub>COOH buffer solution (pKa = 4.757), the amount of H<sup>+</sup> is calculated based on the initial and final pH values, considering the acid-base reaction between CH<sub>3</sub>COOH and H<sub>2</sub>O. Using PhreeqC® hydrogeochemical software and the ThermoddemV1.10_15Dec2020.dat thermodynamic database generated by BRGM (Blanc et al., 2012), the amount of Fe<sup>2+</sup> generated by corrosion reached 5.0 × 10<sup>-2</sup> mol L<sup>-1</sup> (5.0 × 10<sup>-4</sup> mol) when pH value increased from 1 to 4.63; it reached 3.4 × 10<sup>-2</sup> mol L<sup>-1</sup> (3.4 × 10<sup>-4</sup> mol) in the acetate buffer when pH value increased from 4 to 5.41. In the pH range 5.4–10, the decreasing kinetics of AGC-nPd/µZVI induce PFAS-CAP drop. In agreement with the general and basic knowledge on the corrosion of Fe<sup>o</sup> under anoxic conditions (§ 3.2), Fe<sup>2+</sup> remains free in the aqueous phase and at the iron interface in the pH range 0–8.8 at 25 °C. From pH 8.8 to 10.0, Fe(OH)<sub>2</sub> precipitates, in equilibrium with FeOH<sup>+</sup>, onto nPd/µZVI, resulting in drastic decrease in PFAS-CAP.

The sorption mechanism is driven by surface complexation between adsorbed Fe<sup>II</sup> (adsorbed Fe<sup>2+</sup> onto the nPd/µZVI in 4 ≤ pH < 8.8 and adsorbed FeOH<sup>+</sup> onto Fe(OH)<sub>2</sub>-coated nPd/µZVI in 8.8 ≤ pH ≤ 10.0) and the oxygen electron pair of the acetate and sulfonate groups. Surface complexation is aided by electrostatic interactions between adsorbed Fe<sup>II</sup> on the nPd/µZVI hydrophilic solid surface (Peng et al., 2018) and the anionic surfactants, resulting in a first surfactant layer (Fig. 5). A second surfactant layer attaches to the first by their tails, forming a bilayer (or admicelles) (Lawal and Choi, 2018; Hibbs, 2006; Wang and Guo, 2020; Kalam et al., 2021), as illustrated in Fig. 5. The hydrophobic effect in aqueous media causes perfluorinated chains to pack together, enabling dense, ordered packing in micelles (or admicelles) (Krimm, 1980; Ameduri and Boutevin, 2011). According to Wang et al. (Wang et al., 2013), when dissolved, Fe<sup>2+</sup> can complex with the oxygen electron pairs of two surfactant head groups of the bilayer, creating cation
<figure><img src="image_8.png" /><figcaption>**Fig. 4.** Evolution of abatement and desorption of PFOA (a) and PFOS (b) as a function of the initial pH of the aqueous phase. (c) Comparison of the pH values of the aqueous phase before (initial pH) and after each experiment (final pH). (d) Effective influence of pH on PFOA and PFOS removal by nPd/µZVI particles. Conditions: C₀ = 9.29–10 mg L⁻¹, nPd/µZVI = 1 g, T = 90 °C, n = 3.</figcaption></figure>

bridging. The bilayer formation creates a hydrophilic surface, with or without Fe²⁺ bridging. Compared to carboxylate group, the negatively charged structure sulfonate is highly delocalized over the three oxygen atoms, reducing its ability to form strong bonds. The lower complexing capacity of sulfonate groups compared to carboxylate, along with electrostatic attraction and surface complexation, does not explain why

PFOA removal is consistently lower than PFOS across all pH levels. The difference can be attributed to PFOA molecular conformation (Fig. 1): PFOS's sulfonate group aligns with its fluorinated chain, while PFOA's acetate group forms an angle higher than 90°, hindering hydrophobic interactions needed for admicelle formation on the solid surface.

At pH 14, the suspended hydrophobic nPd/µZVI particles oriented

<figure><img src="image_9.png" /><figcaption>**Fig. 5.** Schematic representation of nPd/µZVI/PFAS interaction under acidic and near neutral pH conditions.</figcaption></figure>
the adsorbed fluorocarbon chains outward, indicating that the polar groups of PFAS were directed toward the nPd/µZVI surface. However, at this pH, the excepted increase in reaction rate is responsible for an increase in AGC-nPd/µZVI that leads to the accumulation of FeII as FeO₂⁻²⁻ (Chivot, 2004). The resulting coulombic repulsion between FeO₂⁻²⁻ and the anionic PFAS is expected to hinder PFAS-CAP formation.

In NaOH medium (Wiegmann, 1980), amorphous silica dissolves more readily due to nucleophilic attack of OH⁻ on Si-O-Si bonds, forming soluble silicates. This reaction accelerates with increasing temperature and NaOH concentration. In glasses containing metal oxides such as CaO, Ca²⁺ is released as the glass network dissolves, with NaOH promoting the hydrolysis of Si-O-Ca bonds (Fig. 6) (Li et al., 2022a). According to Peng et al. (Peng et al., 2018), surface Fe(II) species, in the form of adsorbed FeO₂⁻²⁻ and soluble Ca²⁺ interact with silicate anions, in the form of H₂SiO₄²⁻ (Nikolaychuk, 2014), affecting their stability, leading to gel or silicate deposit formation on nPd/µZVI.

The nucleophilic condensation reaction between ≈FeO₂⁻²⁻ and H₂SiO₄²⁻ creates negatively charged Si(IV)-modified nPd/µZVI, where PFAS complex via CaII (Wang et al., 2013), facilitating nPd/µZVI flotation (Fig. 6) (Li et al., 2022a). The increased distance between fluorinated chains prevents hydrophobic interactions between the tails. The same mechanism, in which PFAS anions complex with Ca²⁺ on the SiO₂O₂⁻²⁻ functional group of the vial, is probably also responsible for their sorption onto glass vials.

### 3.3. Insight into PFAS removal at pH 4

#### 3.3.1. Influence of the interaction time

Fig. 7a presents the influence of the interaction time between nPd/µZVI and PFOA or PFOS. Fig. 7b shows the pH values of the aqueous phase over time during interaction of nPd/µZVI with PFOA or PFOS.

Concerning the removal protocol, successive dilutions and UHPLC-MS analyses show the removal efficiency is in the same order of magnitude over the 24 h of interaction (≈ 73 % for PFOA and ≈ 97 % for PFOS). The PFAS sorption equilibrium is reached within 1 h. Under the investigated AGC-nPd/µZVI conditions, the reduction reaction of H⁺/H₂O into H₂ is responsible for the decrease in the buffering capacity of the acetate solution. The pH followed the same trend for the two investigated PFAS. The acetic acid/sodium acetate buffer (pKa 4.76 at 25 °C) maintained its buffering capacity over the 24 h, with pH values exceeding the half-equivalence point by 0.3–0.6 units in the range 6–24 h. This indicates that, under the studied conditions, PFAS do not affect the corrosion process, or they influence it in the same manner. Based on the nPd/µZVI-to-PFAS molar ratio, the corrosion process is independent of the nature of the investigated PFAS.

For the following experiments, the reaction time was set at 3 h such as to maintain pH near the pKa value.

#### 3.3.2. Influence of PFAS-nPd/µZVI ratio

Fig. 8 shows the evolution of PFOA and PFOS removal as a function of the increasing initial PFAS concentration introduced into the slurry containing 1 g of nPd/µZVI.

Within the investigated PFAS concentration range (9 ≤ [PFAS] (mg L⁻¹) ≤ 455), the removal efficiency remained constant at ≈ 72 % for PFOA and ≈ 98 % for PFOS, with the desorbed fraction being of the same order of magnitude.

The sorption isotherms of PFOA and PFOS are presented on Fig. 9. For PFOA, the linear sorption fitted well in the entire concentration domain investigated in this study, with a K_d value of 0.026 L g⁻¹ (26 L kg⁻¹), reflecting moderate sorption. For every 1 kg of nPd/µZVI, the solid holds 26 times more solute than what remains in 1 L of the aqueous phase at equilibrium. PFOA tends to sorb onto the solid, but a significant amount remains dissolved. For PFOS, the linear sorption was accurate for explaining the sorption behavior for [PFOS]₀ < 260 mg L⁻¹, with a K_d value of 0.82 L g⁻¹ (820 L kg⁻¹) in agreement with the high affinity of PFOS for nPd/µZVI. A K_d of 820 L kg⁻¹ for PFOS can explain high

<figure><img src="image_10.png" /><figcaption>Fig. 6. Schematic representation of nPd/µZVI/PFAS at pH 14.</figcaption></figure>

removal efficiencies (≈ 98 %, as observed). The deviation observed for the highest PFOS concentration has not been investigated.

#### 3.3.3. Influence of carbon-fluor chain length on PFSA removal by nPd/µZVI particles

Fig. 10 shows the PFOS (C₈F₁₇SO₃H), PFHxS (C₆F₁₃SO₃H) and PFBS (C₄F₉SO₃H) removal efficiency by Pd/µZVI particles. Based on the error bars, as well as the removal/washing protocol, successive dilutions, and
<figure><img src="image_11.png" /><figcaption>Fig. 7. (a) Influence of the interaction time between nPd/µZVI and PFOA or PFOS. (b) Temporal variation of aqueous-phase pH during experiments involving nPd/µZVI and PFOA or PFOS. Conditions: pH = 4 of the aqueous phase pH before interaction, C₀ = 9.29–10 mg L⁻¹, nPd/µZVI = 1 g, T = 90 °C. n = 3.</figcaption></figure>

<figure><img src="image_12.png" /><figcaption>Fig. 8. Evolution of removal and desorption of PFOA (a) and PFOS (b) as a function of the increasing initial PFAS concentration introduced into the slurry containing 1 g of nPd/µZVI. Conditions: pH = 4, T = 90 °C. n = 3.</figcaption></figure>

UHPLC-MS analyses, sorption appears to be the predominant removal mechanism of PFSAs by nPd/µZVI particles. The results showed a decrease in sorption efficiency with the decrease of the carbon-fluor chain length; 97 % for PFOS, 70 % for PFHxS and 55 % for PFBS. The shortening of the hydrophobic tails progressively prevents the bilayer formation, i.e., the hydrophobic bonding between tails on the hydrophilic surface of nPd/µZVI.

### 3.4. nPd/µZVI recycling at pH 4

Fig. 11 presents the evolution of PFOA and PFOS removal at pH 4 as a function of the number of reuse cycles after methanol washing of nPd/µZVI.

The removal efficiency for PFOA was 15 % lower than that reported in the previous sections, but still within the experimental error range of 10–15 %. The efficiency remained stable for at least 4 cycles, at ≈ 55–70 % for PFOA and ≈ 90–100 % for PFOS. For PFOA, a gradual decrease in efficiency was observed starting from the 5th cycle, reaching ≈ 35–45 % by the 8th cycle. For PFOS, the decrease began from the 6th cycle, dropping to ≈ 75–85 % by the 8th cycle. These results indicate that nPd/

µZVI can be used for at least 4 cycles without significant loss in removal efficiency.

### 3.5. Characterization of nPd/µZVI particles

#### 3.5.1. SEM/EDS analysis

3.5.1.1. Pristine nPd/µZVI. Fig. 12 shows SEM images and EDS analysis of the pristine nPd/µZVI particles, before any interactions. The images reveal microscopic spherical ZVI particles with nanoscale palladium spots, either isolated or coating the ZVI surface. Gold and carbon were introduced during the SEM/EDS sample preparation, where the nPd/µZVI particles were placed on carbon tape and coated with a thin gold layer. The presence of oxygen indicates a thin iron oxide layer on the ZVI. Traces of sodium and silicon likely originated from the biological palladium precipitation process (Rodrigues et al., 2019; Dien et al., 2013).

3.5.1.2. nPd/µZVI particles after interaction with PFOS at pH 14. Fig. 13 shows SEM images and EDS analysis conducted for nPd/µZVI, after
<figure><img src="image_13.png" /><figcaption>Fig. 9. (a) Evolution of the equilibrium PFAS content as a function of the initial PFAS content. (b) PFAS linear sorption isotherms on nPd/µZVI particles. C₀ = 9.29–10 mg L⁻¹, nPd/µZVI = 1 g, T = 90 °C, n = 3.</figcaption></figure>

<figure><img src="image_14.png" /><figcaption>Fig. 10. Comparison between PFOS, PFHxS and PFBS removal by nPd/µZVI. Conditions: pHᵢ = 4, C₀ = 9.29–10 mg L⁻¹, nPd/µZVI = 1 g, T = 90 °C. Error bars correspond to standard deviations (n = 3).</figcaption></figure>

interaction with PFOS at pH 14. Fig. 13a and Fig. 13b show the settled particles, respectively, before and after methanol washing, while Fig. 13c presents the suspended particles. The observation of the settled particles before methanol washing (Fig. 13a) shows ZVI microparticles (Fig. 13a, 1a) and Pd nanoparticle aggregation (Fig. 13a, 3a). A thin film was observed at the surface and between the particles. NaOH crystal, Si, O and Ca as well as potential traces of iron (oxyhydr)oxides were also analyzed (Fig. 13a, 2a, 3a). After the methanol washing (Fig. 13b), nPd/µZVI presented the same characteristics as before methanol washing (Fig. 13a), except for the presence of the thin film between and at the surface of the particles. Although no fluoride was detected by EDS analysis due to the film's thickness (Fig. 13a, 1a, 2a, 3a), it is assumed to be an adsorbed PFOS film, based on methanol's strong efficiency in PFAS desorption. Moreover, compared to Fig. 122a, 2a, 3a, the absence of Ca in Fig. 13b, 2b, 3b supports the PFAS sorption mechanism at pH 14.

The SEM/EDS analysis of the suspended solids after the nPd/µZVI/ PFOS interaction (Fig. 12c) revealed the presence of agglomerated nPd/ µZVI particles at both nanoscopic and microscopic scales, with a diam- eter of 7 µm. Their composition (Fig. 12c, 1c, 2c, 3c, 4c) was similar to that in Fig. 13a. The formation of 7 µm suspended nPd/µZVI particles clearly supports the hydrophobic character of PFAS-Ca-SiO₄-coated nPd/µZVI (Fig. 6). By analogy with the observations and results for

<figure><img src="image_15.png" /><figcaption>Fig. 11. Evolution of PFAS removal as a function of the cycle number of methanol regeneration for (a) PFOA and (b) PFOS. Conditions: pHᵢ = 4, C₀ = 9.29–10 mg L⁻¹, nPd/µZVI = 1 g, T = 90 °C, n = 3.</figcaption></figure>
<figure><img src="image_16.png" /><figcaption>Fig. 12. SEM images and EDS analysis conducted for the pristine nPd/µZVI particles.</figcaption></figure>

PFOA, this also explains the “weak” removal efficiency in PFOA and PFOS at pH 14 because of the loss of adsorbed PFAS onto suspended nPd/µZVI.

3.5.1.3. nPd/µZVI particles after 8 reuse cycles at pH 4. Fig. 14 shows SEM images and EDS analysis conducted for nPd/µZVI after 8 reuse cycles for PFOS and PFOA removal at pH = 4.

The SEM/EDS analysis of nPd/µZVI after 8 reuse cycles (Fig. 11) shows similar results for PFOS and PFOA removal at pH 4, with ZVI exhibiting almost the same size distributions as those observed in Fig. 12 with a thicker oxide coating, Pd aggregates at ≈ 10 µm (EDS analysis 1.a, 1.b and 3.b), and iron (oxyhydr)oxides (EDS analysis 2.a, 2.b and 4.b). The oxidation of ZVI over the reuse cycling is likely due to the washing and drying process under air atmosphere at ambient temperature. Methanol promotes the desorption of PFAS and possibly the removal of the remaining (oxyhydr)oxide shell from the surface of ZVI, exposing the

metal to further oxidation-reduction reactions (Newman, 2008). These reactions contribute to the coverage of the ZVI surface by Fe³⁺ oxidation products under air atmosphere, similar to the oxidation of Fe²⁺ remaining in the 1 mL of methanol from washing, which precipitates as ferrihydrite (Morcos et al., 2022) and then thermodynamically transforms into goethite, magnetite, and lepidocrocite (Das et al., 2011; Cudennec and Lecerf, 2006).

3.5.2. XRD analysis conducted for nPd/µZVI particles after 8 reuse cycles

To complement the SEM/EDS analysis, XRD spectra were obtained to investigate the nature and crystallinity of the different phases constituting the nPd/µZVI solid, including the pristine material before use and after reuse cycles (experiments were conducted after 1 and 8 cycles, Fig. 15). The pristine particles were essentially composed of Fe⁰ (JCPDS card No. 03-065-4899) and Pd⁰ (JCPDS card No. 01-087-0641). After the first cycle, no significant difference with pristine solid was observed.
<figure><img src="image_17.png" /><figcaption>**Fig. 13.** SEM/EDS analysis of nPd/µZVI particles after interaction with PFOS at pH 14: (a) the remaining solid before methanol washing, (b) the remaining solid after methanol washing, and (c) the suspended solid.</figcaption></figure>
<figure><img src="image_18.png" /><figcaption>**Fig. 14.** SEM/EDS analysis of nPd/µZVI particles after 8 recycling cycles for (a) PFOS and (b) PFOA removal.</figcaption></figure>

However, after 8 cycles, the presence of goethite α-FeOOH (JCPDS card No. 01-076-7163) and magnetite Fe₃O₄ (JCPDS card No. 01-088-0315) with traces of maghemite γ-Fe₂O₃ (JCPDS card No. 00-039-1346) was observed, in agreement with the thermodynamic evolution of ferrihydrite into goethite and magnetite (Morcos et al., 2022; Das et al., 2011; Cudennec and Lecerf, 2006). Fe⁰ remained the dominant phase after 8 reuse cycles, consistent with experiments conducted with PFOS, where the removal efficiency remained higher than 80 %. The progressive decrease in removal efficiency is due to the gradual formation of Fe³⁺ oxyhydroxides on nPd/µZVI during washing and drying in air, along with the aggregation of Pd nanoparticles into microparticles. This reduces anoxic galvanic corrosion and decreases Fe²⁺ (pH ≤ 5.41) availability at the nPd/µZVI interface.

<figure><img src="image_19.png" /><figcaption>**Fig. 15.** XRD spectra of the pristine nPd/µZVI, after 1 cycle of PFOA removal and desorption and after 8 cycles of PFOA removal and desorption, with an attribution of the main peaks according to JCPDS cards.</figcaption></figure>
### 3.6. Implications for environmental applications

We disproved any PFAS degradation by Pd/ZVI and confirmed that removal occurred by sorption. This conclusion is supported, first, by the absence of any detectable organic (see Table 3) and inorganic (F⁻) degradation products, up to the analytical limits of our methods and procedures, and second, by the observation of reversible sorption using methanol.

Currently, only hydrothermal alkaline treatments at high temperatures (200–400 °C) (Wu et al., 2019; Li et al., 2022b; Hao et al., 2022) and sonication at very high frequencies (540–1144 kHz) (Awoyemi et al., 2024) have demonstrated effective PFAS degradation. However, these methods are extremely costly. To improve their economic feasibility, PFAS must first be extracted and concentrated. In this context, ZVI-based particles offer a valuable pre-treatment solution by removing PFAS through sorption. Indeed, Zhang et al. demonstrated that nZVI adsorbs PFAS at concentrations ranging from 2.5 µg L⁻¹ to 1 mg L⁻¹ with a maximum sorption capacity of approximately 1000 mg of PFAS per kg of nZVI) (Zhang et al., 2018). Lawal et al. further demonstrated the superior sorption performance of bimetallic Pd/ZVI compared to nZVI and µZVI for PFOA, extending the effective concentration range to 5–80 mg L⁻¹ and achieving a sorption capacity of 16,000 mg of PFOA per kg of Pd/ZVI) (Lawal and Choi, 2018). Our study significantly expands this range to 9.29–455 mg L⁻¹ for both PFOA and PFOS, while still achieving a sorption capacity of 9740 mg of PFOA or PFOS per kg of nPd/µZVI. The nPd/µZVI particles remained effective for at least 4–5 reuse cycles without notable loss of sorption efficiency. For environmental applications, ZVI-based particles can thus be considered as an efficient sorbing material capable of operating effectively over a wide concentration range spanning several orders of magnitude.

Further improvements are needed to enhance sorption capacity in order to compete with granulated activated carbon (GAC) and biochar (Lei et al., 2023; Xiao et al., 2017), which are currently the most commonly used adsorbents for long-chain PFAS removal. For example, unmodified coal-based commercial GAC (F400 grade) has reported maximum Langmuir capacities of 35,693 mg kg⁻¹ for PFOA and 51,138 mg kg⁻¹ for PFOS (Zhang et al., 2021). In comparison, biochar exhibits a broader capacity range, from approximately 1000 to 536,000 mg kg⁻¹ for PFOA and 1000 to over 634,000 mg kg⁻¹ for PFOS (Ramos and Ashworth, 2025; Chavan et al., 2024), depending on the feedstock, preparation method, and surface modification.

Further improvements should also focus on enhancing material reusability, demonstrating process robustness in natural waters, and evaluating the costs associated with: maintaining anoxic conditions to minimize Fe(III) precipitation on the particle surface, thereby reducing the corrosion rate and improving PFAS sorption, keeping the pH below neutrality to maintain a sufficiently high corrosion rate, ensuring the formation of a monolayer on the particle surface with PFAS complexed to Fe²⁺ cations, homogenizing slurries at 90 °C, separating, washing, reusing nPd/µZVI, providing an effective method for PFAS removal for subsequent defluorination, and scale-up and implementation under field conditions. Indeed, in terms of Technology Readiness Level (TRL), GAC technology has already been deployed at large scales for water treatment (Ross et al., 2018).

Regarding the cost of raw materials, the price of GAC can range from $0.15 to $2.90 per kg, (Stefaniuk et al., 2016). In comparison, biochar is typically 3 to 6 times less expensive (Deng et al., 2015) while Fe-based particles range from $1.20 to $337 per kg (Stefaniuk et al., 2016). It is, however, difficult to provide an exact cost because it is highly dependent on the synthesis method, including the addition of nobler metals such as palladium, which increases the price of the particles, as well as the desired particle size (Visentin et al., 2019).

Practically, as the particle size of ZVI in nPd/µZVI is around 2 µm, the sorption rate (1 to 3 h) is relatively high compared to those reported for carbon-based sorbents (≤ 40 h), whose particle size is usually greater than 100 µm (Zhang et al., 2016). Another advantage is the simple

regeneration of the used materials, with only 3 methanol washings, without significant loss of removal efficiency for several cycles, compared to thermal regeneration under high (400–800 °C) or very high temperature (1000–1200 °C), supercritical CO₂ extraction, or other chemical regeneration, using acidic and/or alkaline aqueous solutions as well as organic/inorganic mixed solutions, which allow the particles to be regenerated 4 to 5 times, up to 10 times, with diminishing returns in terms of efficiency (Deng et al., 2015; Didenko et al., 2025; Chulauangaksorn et al., 2014; Du et al., 2014; Larasati et al., 2020; Matthis and Carr, 2018; McNamara et al., 2018). Hence, this approach may reduce the need for frequent addition of fresh ZVI-based particles to maintain high PFAS removal efficiency, thereby enhancing economic competitiveness.

While the aim is to develop a competitive sorption process, further technical and economic evaluations are required, in addition to process optimization.

### 4. Conclusion

This work provides an in-depth investigation of PFAS-nPd/µZVI interactions, focusing on the mechanism, performance, and nPd/µZVI reuse potential. Experiments were conducted in sealed batches using ultrasonically homogenized slurries, with 1 g of nPd/µZVI and PFAS concentrations ranging from 9.29 to 485 mg L⁻¹ (including PFOA and PFOS) across a pH range of 1 to 14. PFHxS and PFBS were also used to analyze the influence of the carbon-fluorine chain length.

PFAS sorption is the predominant abatement mechanism. It is driven by the anoxic galvanic corrosion of nPd/µZVI (AGC-nPd/µZVI), which modifies the physical and chemical conditions of the liquid phase and the surface chemistry of nPd/µZVI as a function of pH.

In the pH range 4–5.4, massive production of adsorbed Fe²⁺ is responsible for PFAS complexation adsorption process (PFAS-CAP) onto nPd/µZVI. In the pH range 5.4–10, the decreasing kinetics of AGC-nPd/ µZVI induce PFAS-CAP drop. From pH 8.8 to 10.0, Fe(OH)₂ precipitates (in equilibrium with FeOH⁺) onto nPd/µZVI, resulting in drastic decrease in PFAS-CAP.

The sorption mechanism is driven by surface complexation between adsorbed Fe¹¹ (adsorbed Fe²⁺ onto the nPd/µZVI in 4 ≤ pH < 8.8 and adsorbed FeOH⁺ onto Fe(OH)₂-coated nPd/µZVI in 8.8 ≤ pH ≤ 10.0) and the oxygen electron pair of the acetate and sulfonate groups. Surface complexation is facilitated by electrostatic interactions between adsorbed Fe¹¹ on the nPd/µZVI hydrophilic solid surface and the anionic surfactants, resulting in a first surfactant layer. A second layer of surfactants attaches to the first through hydrophobic interactions between the tails, forming a bilayer that is stabilized by fluorine-fluorine van der Waals interactions between the tails and by Fe²⁺, which complexes with the oxygen electron pairs of the two surfactant head groups in the bilayer when Fe²⁺ predominates.

We assume that the difference in removal efficiency between PFOA (70–80 %) and PFOS (90–100 %) results from steric hindrance: the angle greater than 90° formed by the acetate group of PFOA hinders the hydrophobic interactions necessary for admicelle formation and consolidation on the solid surface. Additionally, the shortening of the hydrophobic tails progressively prevents bilayer formation.

At pH 14, NaOH induces the leaching of silicates and Ca²⁺, which interact to form negatively charged SiO₂-modified nPd/µZVI, where PFAS complex via Ca²⁺, facilitating nPd/µZVI flotation. The increased distance between the fluorinated chains hinders hydrophobic interactions between the tails.

Under the investigated conditions, a linear sorption model accurately described PFOS removal at pH 4. nPd/µZVI is as effective as activated carbon for the efficient removal of long-chain PFAS.

At pH 4, nPd/µZVI were demonstrated to be reusable for at least 4 to 5 cycles without significant loss of sorption efficiency. The progressive decrease in removal efficiency is due to the gradual formation of Fe¹¹¹ oxyhydroxides on nPd/µZVI during washing and drying in air, along
with the aggregation of Pd nanoparticles into microparticles. This re-
duces anoxic galvanic corrosion and Fe²⁺ availability at the nPd/µZVI
interface. The nPd/µZVI washing protocol under air atmosphere is still
being optimized. Ultimately, nPd/µZVI may be of particular interest in a
treatment train for PFAS pre-concentration, where the use of an organic
solvent such as methanol could both extract and concentrate PFAS while
regenerating the particles.

CRediT authorship contribution statement

**Raphaël Tur:** Writing – review & editing, Writing – original draft, Methodology, Investigation, Formal analysis, Conceptualization. **Stéphanie Betelu:** Writing – review & editing, Writing – original draft, Supervision, Methodology, Investigation, Formal analysis, Conceptualization. **Romain Rodrigues:** Writing – review & editing, Writing – original draft, Methodology, Investigation, Formal analysis, Conceptualization. **Stéfan Colombano:** Supervision, Funding acquisition, Conceptualization. **Dorian Davarzani:** Supervision. **Sébastien Bristeau:** Writing – review & editing, Formal analysis. **Julien Grandclément:** Supervision. **Arnaud Perrault:** Funding acquisition. **Julie Lions:** Project administration, Funding acquisition. **Eric D. van Hull ebusch:** Writing – review & editing, Supervision, Funding acquisition. **Ioannis Ignatiadis:** Writing – review & editing, Writing – original draft, Supervision, Methodology, Investigation, 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

PROMISCES project has received funding from the European Union's Horizon 2020 research and innovation program under grant agreement N° 101036449. The Doctoral thesis of Raphaël Tur was funded by the French National Research and Technology Agency (ANRT, CIFRE N° 21/1912) and the private company COLAS Environment. SEM/EDS analyses conducted by Stephan Borensztajn were supported by the IPGP multidisciplinary program PARI, and by Paris-IdF region SESAME Grant no. 12015908. The authors acknowledge Nicolas Maubec and Raphaël Dupuy for XRD measurements and refinement, as well as Flora Soler, Aya Messaoudi, Emeline Coisy and Stephane Touzelet for their help in the experimental work.

Data availability

Data will be made available on request.

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