# Molybdenum and Antimony immobilization by iron oxalate salt in contaminated soils

Amandine Dumas, Laurent de Windt, Ivan Serclerat, David S Kosson, Bruno Huet, Florence Sanchez, Nouffou Tapsoba

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Amandine Dumas, Laurent de Windt, Ivan Serclerat, David S Kosson, Bruno Huet, et al., Molybdenum and Antimony immobilization by iron oxalate salt in contaminated soils. Applied Geochemistry, 2025, 195, pp.106626. 10.1016/j.apgeochem.2025.106626. hal-05400839

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# Molybdenum and Antimony immobilization by iron oxalate salt in contaminated soils

Amandine Dumas<sup>a,b</sup>, Laurent De Windt<sup>b,*</sup>, Ivan Serclerat<sup>c</sup>, David S. Kosson<sup>d</sup>, Bruno Huet<sup>a</sup>, Florence Sanchez<sup>d</sup>, Nouffou Tapsoba<sup>a</sup>

## ARTICLE INFO

Editorial Handling by: Dr C S Eckley

**Keywords:**

Iron salt

Metal oxyanion

Modeling

Leaching assessment

S/S

Stabilization

## ABSTRACT

This study introduces iron oxalate dihydrate (FeOx) as an innovative and highly effective treatment for the chemical stabilization of trace levels of Mo and Sb in soils. The effectiveness of the FeOx treatment was evaluated for a Mo-contaminated limestone tunnel sludge (LTS) and Mo/Sb-contaminated industrial wasteland soil (IWS) with organic matter. FeOx reactivity and oxyanion retention were investigated using TGA, XRD, optical microscopy and standardized US/EU leaching tests. Mineralogical changes and the pH-dependence on Mo sorption were further analyzed by geochemical modeling. Under near-neutral pH and oxidizing conditions, FeOx dissociates, releasing Fe²⁺ and C₂O₄²⁻ ions. The Fe²⁺ is subsequently oxidized to Fe³⁺, forming poorly crystallized hydrous ferric oxides (HFOs) that are uniformly distributed. This process induces acidification, partially dissolving calcite and releasing Ca²⁺, which reacts with C₂O₄²⁻ to form calcium oxalate monohydrate (whewellite). Mo and Sb immobilization occurs through sorption onto HFOs, achieving effective retention at ≥1 wt% FeOx for LTS and ≥5 wt% for IWS, meeting European inert waste standards. The fast transformation of FeOx into HFOs and whewellite enables Mo and Sb retention within 6 h. Overall, leaching decreased by 67 % for Mo in LTS with 2 wt% FeOx, and by 88 % for Sb and 90 % for Mo in IWS with 6 wt% FeOx. Retention stability was observed at pH ≤ 8, with no significant competition with sulfate or acetate ions. These results highlight FeOx as a promising solution for industrial applications requiring rapid chemical stabilization of contaminants. However, further research is needed to assess long-term stability under variable redox conditions.

## 1. Introduction

Antimony (Sb) and molybdenum (Mo) are trace elements of environmental concern due to their toxicity and increased presence in the environment as a result of industrial activities such as metallurgy and plastic production (Pradhan et al., 2023; Smedley and Kinniburgh, 2017; Xu et al., 2013). While Mo is an essential nutrient at low concentrations, elevated levels can lead to adverse health effects including anemia, liver dysfunction, and osteoporosis (Smedley and Kinniburgh, 2017). Similarly, Sb is considered a priority pollutant by the European Union and the U.S. Environmental Protection Agency due to its high toxicity and bioaccumulation potential (Pradhan et al., 2023). Molybdate (MoO₄²⁻) and antimonate (Sb(OH)₆) are highly mobile oxyanions, making them challenging to immobilize and complicating their

management in contaminated environments (Tang et al., 2023).

Soil stabilization through chemical treatment is among the most cost-effective and scalable methods for managing trace contaminants in industrial contexts (Rahman et al., 2023; Zhang et al., 2021). However, most research on chemical stabilization has focused on cationic metals, leaving significant knowledge gaps regarding oxyanions, particularly MoO₄²⁻, which remains largely understudied (Bao et al., 2022; Chen et al., 2023; Kumar et al., 2024; Rahman et al., 2023; Tang et al., 2023; Zhang et al., 2021). Addressing this gap requires studying the behavior of Sb and Mo oxyanions in realistic soils that exhibit complex mineralogy, multiple pollutant interactions, and low contaminant concentrations on the order of a few mg/kg (Dumas et al., 2025; Smedley and Kinniburgh, 2017; Tang et al., 2023). Dissolved organic matter (OM) and oxyanions such as sulfates are common in waste materials and can

<footnote><sup>a</sup> Holcim Innovation Center, St Quentin Fallavier, France</footnote>

<footnote><sup>b</sup> Mines Paris, PSL University, Centre for Geosciences and Geo-engineering, Fontainebleau, France</footnote>

<footnote><sup>c</sup> Geocycle, St Quentin Fallavier, France</footnote>

<footnote><sup>d</sup> Civil and Environmental Engineering Department, Vanderbilt University, Nashville, USA</footnote>

<footnote>* Corresponding author.</footnote>

<footnote>E-mail address: laurent.de windt@mines-paristech.fr (L. De Windt).</footnote>
compete with Sb and Mo for sorption sites, a factor that has to be considered in realistic remediation scenarios (Goldberg and Forster, 1998; Han et al., 2020; Kumar et al., 2024; Peng et al., 2023; Xu et al., 2013).

Iron (hydr-)oxides play a key role in the retention of pollutants due to their ability to sorb various contaminants, including Mo and Sb oxyanions (Peng et al., 2023; Yang and Wang, 2021). This sorption capacity has made iron (hydr-)oxides a common choice for soil and water treatment (Gong et al., 2022; Haleemat Iyabode Adegoke et al., 2013; Shi et al., 2021). However, their efficiency is highly pH-dependent; above the point of zero charge (pzc), the negatively charged surfaces are less effective at retaining anions (Haleemat Iyabode Adegoke et al., 2013; Han et al., 2020; Kalaitzidou et al., 2019).

Recently, the use of iron salts for oxyanion immobilization in soil has attracted attention, although it remains less studied than iron (hydr-) oxides or zero-valent iron (Tian et al., 2024; Wang et al., 2021). Iron salts facilitate a uniform dispersion of iron in the soil and can dissociate to form poorly crystallized, high-surface-area, hydrous ferric oxides (HFOs). However, conventional salts like iron chloride and sulfate introduce undesirable counterions, which can be environmentally problematic (Brandely et al., 2022b; Van Caneghem et al., 2016). This study proposes the use of iron oxalate dihydrate (FeOx) as an innovative and environmentally friendly stabilizing agent for Sb and Mo immobilization (Tang et al., 2023). To our knowledge, FeOx has not previously been investigated for soil stabilization. Oxalate biodegradability and its ability to form highly insoluble complexes with common soil cations such as calcium and magnesium, makes it particularly suitable for this application (Baran and Monje, 2008; Uren, 2018). Beyond sorption, the precipitation of Sb and Mo minerals such as FeSbO₄, Mo₃O₈ or CaMoO₄ (powellite) may contribute to the immobilization of Sb and Mo when present at high concentrations (Mancini et al., 2020; Van Caneghem et al., 2016).

This study aims to bridge the knowledge gap by investigating the immobilization of Sb and Mo oxyanions through iron oxalate treatment in two distinct excavated contaminated geomaterials. For clarity, both materials are referred to as “soils” in this article. The first soil is a limestone tunnel sludge (LTS). It is primarily composed of calcite and dolomite and exhibits geogenic Mo contamination. The second soil is an industrial wasteland soil (IWS). It is predominantly composed of aluminosilicates and contaminated with both Sb and Mo, while also being characterized by high OM content. A combined approach using thermogravimetric analysis (TGA), X-ray diffraction (XRD), and geochemical modeling was employed to investigate retention mechanisms and mineralogical transformations. Standard leaching tests were conducted to assess chemical durability and immobilization effectiveness. US EPA 1313 was applied to evaluate pH sensitivity across a range of 2–13, while EN 12457-2 was used to assess immobilization over time, with stabilization periods ranging from 6 h to 90 days. Additionally, US EPA 1311 (TCLP) was also used as a standard regulatory indicator to assess treatment efficiency.

This study aims to highlight the potential of Mo and Sb immobilization by FeOx as a rapid and sustainable solution for industrial soil remediation applications.

## 2. Materials and methods

### 2.1. Contaminated soils

The first material is a limestone tunnel sludge (LTS) produced during the “Grand Paris Express” subway extension project in Paris, France. A picture of the LTS material can be found in S.I. (Fig. S1–A). After extraction by a tunnel-boring machine, the sludge underwent multiple processing stages. First, lime was incorporated to lower the water content and enhance the handling characteristics of the material. The material was subsequently stockpiled for a prolonged period. For the purposes of this study, a 600 kg sample was collected, air-dried, ground

to <4 mm, and homogenized by quartering. These preliminary processes likely led to oxidizing conditions. Qualitative XRD analysis indicated that the material primarily consists of dolomite, calcite and quartz (Table 1). Gypsum was also detected, and its relatively high solubility makes it a potential source of sulfate release. The organic carbon content of the material is low (0.4 wt%). Molybdenum is present in trace amounts at 7 mg/kg of dry matter (DM). Mo(VI) as molybdate MoO₄⁻ is likely the main species under these oxidizing conditions (Smedley and Kinniburgh, 2017; Xu et al., 2013). BET analysis indicated a specific surface area of 12 m²/g, reflecting the high fineness of the material.

The second material is an industrial wasteland soil (IWS) from France, with limited historical data on its geographic origin or the type of industry (Fig. S1–B). The material was stored in a pile and periodically turned over for at least one year. Preparation for this study involved sieving to remove particles greater than 4 mm, resulting in a 13 % oversize fraction primarily composed of rocks, roots, wood pieces, and plastic waste. A picture of the oversize fraction is presented in Fig. S2. The sieved soil was then homogenized by quartering. As with the first soil, these preliminary processing steps likely created oxidizing conditions. According to XRD analysis, this soil is primarily composed of silicate minerals, particularly feldspars (Table 1). It has a high organic carbon content of 4.4 wt%. Mo and Sb are present at concentrations of 3 mg/kg and 15 mg/kg DM, respectively. Under oxidizing conditions, Sb is expected to exist mainly as Sb(V), antimonate Sb(OH)₆, though Sb(III) may also coexist (Tang et al., 2023; Zhang et al., 2021). BET analysis indicated a specific surface area of 4 m²/g, lower than that of the LTS soil but still relatively high due to the significant fraction of fine particles. Overall, IWS presents multi-oxyanion contamination, a complex mineralogical matrix, and a high organic matter content.

In summary, the two soils have distinct contamination origins: the first one is associated with geogenic contamination, while the second one is likely of anthropogenic origin. Their characteristics differ notably in mineralogical composition (mainly carbonates in the first vs. feldspars in the second) and organic matter content (low in the first and high in the second). Both soils contain trace levels of contaminants, with oxidizing conditions favoring Mo(VI) and Sb(V) as the predominant species.

Additional data on the elemental composition, including oxides and trace elements, are provided in S.I. (Tables S1 and S2).

### 2.2. Treatment protocol

The iron II oxalate dihydrate FeC₂O₄·2H₂O (FeOx), reagent grade

<table><caption>**Table 1**
Qualitative mineralogical composition of the limestone tunnel sludge (LTS) and industrial wasteland soil (IWS), as estimated from XRD analysis.</caption><thead><tr><th rowspan="2">Minerals</th><th rowspan="2">Formula</th><th colspan="2">Occurrence</th></tr><tr><th>LTS</th><th>IWS</th></tr></thead><tbody><tr><td colspan="4"><strong>Carbonates</strong></td></tr><tr><td>Calcite</td><td>CaCO<sub>3</sub></td><td>+++</td><td>+</td></tr><tr><td>Dolomite</td><td>CaMg(CO<sub>3</sub>)<sub>2</sub></td><td>+++</td><td></td></tr><tr><td colspan="4"><strong>Silicates</strong></td></tr><tr><td>Anorthite</td><td>CaAl<sub>2</sub>Si<sub>2</sub>O<sub>8</sub></td><td>+</td><td></td></tr><tr><td>Falcondoite</td><td>(Ni,Mg)<sub>4</sub>Si<sub>6</sub>O<sub>15</sub>(OH)<sub>2</sub>·6H<sub>2</sub>O</td><td>+</td><td></td></tr><tr><td>Palygorskite</td><td>Mg<sub>5</sub>(Si,Al)<sub>8</sub>O<sub>20</sub>(OH)<sub>2</sub>·8H<sub>2</sub>O</td><td>+</td><td></td></tr><tr><td>Quartz</td><td>SiO<sub>2</sub></td><td>++</td><td>+++</td></tr><tr><td>Albite</td><td>NaAlSi<sub>3</sub>O<sub>8</sub></td><td></td><td>++</td></tr><tr><td>Clinochlore</td><td>(Mg,Fe,Al)<sub>6</sub>(Si,Al)<sub>4</sub>O<sub>10</sub>(OH)<sub>8</sub></td><td></td><td>+</td></tr><tr><td>Kaolinite</td><td>Al<sub>2</sub>Si<sub>2</sub>O<sub>5</sub>(OH)<sub>4</sub></td><td></td><td>++</td></tr><tr><td>Microcline</td><td>KAlSi<sub>3</sub>O<sub>8</sub></td><td></td><td>++</td></tr><tr><td>Muscovite</td><td>KAl<sub>2</sub>(AlSi<sub>3</sub>O<sub>10</sub>)(OH)<sub>2</sub></td><td></td><td>++</td></tr><tr><td colspan="4"><strong>Sulfate minerals</strong></td></tr><tr><td>Celestite</td><td>SrSO<sub>4</sub></td><td>+</td><td></td></tr><tr><td>Gypsum</td><td>CaSO<sub>4</sub>·2H<sub>2</sub>O</td><td>+</td><td></td></tr><tr><td colspan="4"><strong>Iron (hydr-)oxides</strong></td></tr><tr><td>Hematite/Goethite</td><td>Fe<sub>2</sub>O<sub>3</sub>/FeO(OH)</td><td>+</td><td>+</td></tr></tbody></table>
(>99.7 %), was supplied by Sigma Aldrich in the form of a fine powder (<60 µm). It consists of humboldtine according to the XRD spectra and the derivative weight curve (TGA) of FeOx provided in S.I. (Fig. S3 and Fig. S4).

The treatment protocol was performed by mechanically mixing the soil with the treatment for 3 min. During the stabilization period, defined as the interval between the chemical agent addition and the leaching test, samples were stored at 22 ± 1 °C in sealed plastic bags to minimize interaction with air and water loss.

To evaluate the efficiency of FeOx and determine the appropriate treatment dosage to remain below the European threshold limits (0.5 mg/kg DM for Mo and 0.06 mg/kg DM for Sb (OJEU, 1999)), several quantities were tested on both soils: 0.5, 1, 2, 4, 5, 6, 8 wt%. The leaching behavior after treatment addition was assessed according to the European batch leaching standard EN 12457-2 as mandated by European regulation (AFNOR, 2002). A stabilization period of 7 days was selected as a compromise, providing sufficient time for the immobilization of Mo and Sb while meeting the requirements for rapid industrial application (Brandely et al., 2022b).

Following the determination of the optimal FeOx dosage for immobilization of Mo and Sb in the contaminated soils, a detailed assessment of the leaching behavior of this selected formulation was conducted using standardized protocols. To evaluate the kinetics of immobilization, samples were collected at various stabilization times and subjected to two regulatory leaching tests: the European EN 12457-2 (AFNOR, 2002) and the U.S. EPA Method 1311 (TCLP) (US EPA, 1992). The EN 12457-2 test, designed for granular materials, was performed after stabilization periods of 0.25, 1, 2, 7, 28, 56, and 90 days, using a liquid-to-solid (L/S) ratio of 10 L/kg (DI water) and a contact time of 24 h. In parallel, the US EPA 1311 test, which involves leaching with dilute acetic acid, was applied after 1, 7, and 28 days of stabilization, using an L/S ratio of 20 L/kg and an 18-h contact time. To complement these assessments, the US EPA Method 1313 (US EPA, 2012) was conducted after 28 days of stabilization, once the mineralogical assemblage of the treated soils had stabilized (as discussed in Section 3.3). This test, conducted at an L/S ratio of 10 L/kg with a contact time of 24 h, applies nine target pH values ranging from 2 to 13 (adjusted using HNO₃ or NaOH in deionized water) to evaluate the influence of pH on leaching behavior. All the leaching tests followed standardized batch procedures, and further details on the experimental protocols can be found in the referenced methodologies. For comparative purposes, the same suite of leaching tests was also applied to the untreated soils.

## 2.3. Solid phases analysis

Two complementary analytical methods were employed to characterize the elemental composition of the untreated soils: wavelength-dispersive X-ray fluorescence (XRF) spectroscopy, performed with an S8 TIGER Series 2 spectrometer (Bruker), and acid digestion using aqua regia in combination with microwave-assisted mineralization (ISO 54321:2020), followed by elemental analysis via inductively coupled plasma optical emission spectrometry (ICP-OES) according to ISO 11885:2009, using a PerkinElmer Avio 500DV instrument. The total organic carbon (TOC) content was derived as the difference between total carbon (TC) and total inorganic carbon (TIC). TC was determined by high-temperature combustion of the sample under an oxygen atmosphere, reaching up to 1000 °C, whereas TIC was assessed by thermal decomposition of carbonates at the same temperature under nitrogen flow. The CO₂ generated during each analysis was measured using infrared (IR) detection with an RC612 C/H analyzer (LECO). Additionally, the mineralogical profile of the soils (dried samples powdered to <60 µm) was qualitatively examined by X-ray diffraction (XRD) on a Bruker D8 Advance diffractometer (Cu-Kα radiation source, operating at 40 mA and 25 kV), employing a continuous scan from 3° to 65°, completed in 10 min over 4336 steps. Thermogravimetric analysis (TGA) was carried out on a TGA/DSC 3+ (Mettler) using 50 mg of

sample, with a resolution of 1 µg, under an inert nitrogen atmosphere from 30 °C to 600 °C at a heating rate of 2.5 °C/min in aluminum crucibles. The specific surface area of the soils was determined by nitrogen adsorption according to the Brunauer–Emmett–Teller (BET) theory, using a Tristar II 3020 analyzer (Micromeritics). Optical microscopy observations were conducted using a Zeiss Axiovert 200 M at a 20x magnification.

For the treated samples, TGA and XRD analyses were specifically conducted at the highest tested dosage (8 wt%) to maximize the detection of possible phase transformations.

## 2.4. Eluates analysis

The final pH and redox potential (Eh) were measured prior to filtration (0.45 µm, PVDF). The pH was measured using a SenTix 41 probe from WTW. The redox potential (Eh) was calculated using the equation: $E_h = U_m + U_{ref}$ where Eh represents the redox potential relative to the standard hydrogen electrode, $U_m$ is the redox potential reading measured with a SenTix® ORP-T 900(-P) probe (WTW), and $U_{ref}$ is the reference potential (207 mV at 25 °C). The concentrations of dissolved organic carbon (DOC) were determined via total carbon analysis in accordance with standard EN 1484, using a TOC-L CPH/CPN analyzer from SHIMADZU. Metal and metalloid contents were quantified by inductively coupled plasma mass spectrometry/optical emission spectrometry (ICP-MS/OES) following the EN ISO 11885:2009 protocol on a PerkinElmer Avio 500DV instrument. Anion concentrations were analyzed via ion chromatography, following the ISO 15923-1:2013 standard, using a Thermo Scientific Dionex ICS-6000 HPIC system.

## 2.5. Geochemical modeling

The geochemical module CHESS, integrated within the HYTEC code (van der Lee et al., 2003) was employed to simulate the mineralogical assemblages and to gain insights into the retention mechanisms of $MoO_4^{2-}$ in the LTS case. All chemical reactions (aqueous chemistry, sorption and dissolution/precipitation of minerals) were modeled at thermodynamic equilibrium at 25 °C. The B-dot model was applied to account for activity corrections in solution. The thermodynamic constants were obtained from the ThermoChimie database, version 11 (Giffaut et al., 2014). Two additional sets of data for Fe and Mg oxalate salts were taken from the literature. Table 2 presents the formation constants of the minerals included in the model. The initial states of the untreated LTS and after addition of FeOx were calculated under equilibrium with atmospheric $CO_2$ and $O_2$.

The sorption of $MoO_4^{2-}$ and $SO_4^{2-}$ onto HFOs was modeled using the double-layer surface complexation dataset of Dzombak and Morel (Table 3) (Dzombak and Morel, 1990). The specific surface area of HFOs was set to 240 m²/g, consistent with the surface area of natural ferrihydrite precipitates (Filip et al., 2007; Weidler, 1997). For $MoO_4^{2-}$, Dzombak and Morel did not fit any experimental data sets but used an extrapolation method to estimate the constant. Based on adjustments previously performed by Brandely (Brandely et al., 2022a, 2025), the logK of the predominant sorbed species $\equiv S^{S,w}OHMoO_4^{2-}$ had to be increased from 2.4 to 4.1 to properly fit the present experimental data (Fig. S5). Gustafsson experimentally determined an intermediate logK value (3.15) for the same surface reaction (Gustafsson, 2003). Only sulfate competitive sorption on HFOs was considered due to its high concentration in leachates (Table S3). Other oxyanions, whether metalloids or not, were either absent or present only in trace amounts in the total content analysis of LTS (Table S1) and were not detected in the leaching eluate analyses (Table S3), particularly phosphate, which is a strong competitor for sorption onto HFOs (Han et al., 2020; Wang et al., 2022b). Oxalates were also excluded as potential competing anions for HFO sorption sites, as they were immobilized in the form of calcium oxalate (Sec.3.1.2) and thus present at low concentration in the solution (Table S3).
<table><caption>**Table 2**

Thermodynamic formation constants of the minerals included in the geochemical model.</caption><thead><tr><th>Mineral</th><th>Formation reaction</th><th>LogK (25 °C)</th><th>Reference</th></tr></thead><tbody><tr><td>Calcite</td><td>Ca<sup>2+</sup> + HCO<sub>3</sub><sup>-</sup> ⇌ CaCO<sub>3</sub> + H<sup>+</sup></td><td>8.48</td><td>Giffaut et al. (2014)</td></tr><tr><td>Dolomite</td><td>Ca<sup>2+</sup> + Mg<sup>2+</sup> + HCO<sub>3</sub><sup>-</sup> ⇌ CaMg(CO<sub>3</sub>)<sub>2</sub></td><td>17.13</td><td>Giffaut et al. (2014)</td></tr><tr><td>Gypsum</td><td>Ca<sup>2+</sup> + SO<sub>4</sub><sup>2-</sup> + 2H<sub>2</sub>O ⇌ CaSO<sub>4</sub>·2H<sub>2</sub>O</td><td>4.61</td><td>Giffaut et al. (2014)</td></tr><tr><td>Celestite</td><td>Sr<sup>2+</sup> + SO<sub>4</sub><sup>2-</sup> ⇌ SrSO<sub>4</sub></td><td>6.62</td><td>Giffaut et al. (2014)</td></tr><tr><td>Quartz</td><td>H<sub>4</sub>SiO<sub>4</sub> ⇌ SiO<sub>2</sub> + 2H<sub>2</sub>O</td><td>3.74</td><td>Giffaut et al. (2014)</td></tr><tr><td>Anorthite</td><td>2Al<sup>3+</sup> + Ca<sup>2+</sup> + 2H<sub>4</sub>SiO<sub>4</sub> ⇌ Al<sub>4</sub>Ca(SiO<sub>4</sub>)<sub>4</sub> + 8H<sup>+</sup></td><td>-25.31</td><td>Giffaut et al. (2014)</td></tr><tr><td>Ferrihydrite (amorphous)</td><td>Fe<sup>3+</sup> + 3H<sub>2</sub>O ⇌ Fe(OH)<sub>3</sub> + 3H<sup>+</sup></td><td>-3.92</td><td>Giffaut et al. (2014)</td></tr><tr><td>Humboldtine</td><td>Fe<sup>2+</sup> + C<sub>2</sub>O<sub>4</sub><sup>2-</sup> + 2H<sub>2</sub>O ⇌ FeC<sub>2</sub>O<sub>4</sub>·2H<sub>2</sub>O</td><td>6.50</td><td>De Gregorio et al. (2023)</td></tr><tr><td>Whewellite</td><td>Ca<sup>2+</sup> + C<sub>2</sub>O<sub>4</sub><sup>2-</sup> + H<sub>2</sub>O ⇌ CaC<sub>2</sub>O<sub>4</sub>·H<sub>2</sub>O</td><td>8.73</td><td>Giffaut et al. (2014)</td></tr><tr><td>Weddellite</td><td>Ca<sup>2+</sup> + C<sub>2</sub>O<sub>4</sub><sup>2-</sup> + 2H<sub>2</sub>O ⇌ CaC<sub>2</sub>O<sub>4</sub>·2H<sub>2</sub>O</td><td>8.30</td><td>Giffaut et al. (2014)</td></tr><tr><td>Magnesium Oxalate</td><td>Mg<sup>2+</sup> + C<sub>2</sub>O<sub>4</sub><sup>2-</sup> + H<sub>2</sub>O ⇌ MgC<sub>2</sub>O<sub>4</sub>·H<sub>2</sub>O</td><td>5.68</td><td>Parkhurst and Appelo (1999)</td></tr><tr><td>Powellite</td><td>Ca<sup>2+</sup> + MoO<sub>4</sub><sup>2-</sup> ⇌ CaMoO<sub>4</sub></td><td>7.90</td><td>Giffaut et al. (2014)</td></tr></tbody></table>

<table><caption>**Table 3**

Surface complexation constants of sorption on HFOs included in the model; the site densities of the strong sites ≡ S<sup>s,w</sup>OH and weak sites ≡ S<sup>s,w</sup>OH are 0.09 and 3.75 µmol/m<sup>2</sup>, respectively.</caption><thead><tr><th>Reactions of sorption</th><th>logK (25 °C)</th><th>Reference</th></tr></thead><tbody><tr><td>≡ S<sup>s,w</sup>OH ⇌ ≡ S<sup>s,w</sup>O<sup>-</sup> + H<sup>+</sup></td><td>-8.9</td><td>Dzombak and Morel (1990)</td></tr><tr><td>≡ S<sup>s,w</sup>OH + H<sup>+</sup> ⇌ ≡ S<sup>s,w</sup>OH<sub>2</sub><sup>+</sup></td><td>7.3</td><td>Dzombak and Morel (1990)</td></tr><tr><td>≡ S<sup>s,w</sup>OH + SO<sub>4</sub><sup>2-</sup> ⇌ ≡ S<sup>s,w</sup>OHSO<sub>4</sub><sup>2-</sup></td><td>0.8</td><td>Dzombak and Morel (1990)</td></tr><tr><td>≡ S<sup>s,w</sup>OH + SO<sub>4</sub><sup>2-</sup> + H<sup>+</sup> ⇌ ≡ S<sup>s,w</sup>SO<sub>4</sub> + H<sub>2</sub>O</td><td>7.8</td><td>Dzombak and Morel (1990)</td></tr><tr><td>≡ S<sup>s,w</sup>OH + MoO<sub>4</sub><sup>2-</sup> ⇌ ≡ S<sup>s,w</sup>OHMoO<sub>4</sub><sup>2-</sup></td><td>4.1</td><td>Brandely et al. (2025)</td></tr><tr><td>≡ S<sup>s,w</sup>OH + MoO<sub>4</sub><sup>2-</sup> + 2H<sup>+</sup> + H<sub>2</sub>O ⇌ ≡ S<sup>s,w</sup>OMo(OH)<sub>5</sub></td><td>17.9</td><td>Dzombak and Morel (1990)</td></tr></tbody></table>

To validate our modeling approach, we first conducted preliminary experiments to determine the time required to reach equilibrium between soil and leachate - a critical prerequisite for thermodynamic modeling. As shown in Fig. S6, while the Limestone Tunnel Sludge (LTS) reached equilibrium within 24 h, consistent with standard leaching protocol durations, the IWS required a substantially longer period. This extended equilibration time demonstrates that the assumption of thermodynamic equilibrium is not satisfied for IWS under standard leaching conditions, justifying its exclusion from the modeling. The slower geochemical reactivity of IWS would require the development of more

sophisticated kinetic models. In addition, the high organic matter content in IWS introduces further modeling challenges, necessitating the inclusion of specific organic speciation and complexation processes. Given these complexities, simulating the behavior of IWS would demand a dedicated study employing more advanced modeling approaches and was therefore beyond the scope of the present work.

### 3. Results and discussion

#### 3.1. Effects of FeC₂O₄·2H₂O addition on mineralogical assemblage

##### 3.1.1. Hydrous ferric oxide precipitation

Upon contact with water, iron oxalate dissociates into Fe<sup>2+</sup> and C₂O₄²⁻ ions, as follows:

$$FeC_{2}O_{4} \cdot 2H_{2}O(s) \rightarrow Fe^{2+} + C_{2}O_{4}^{2-} + 2H_{2}O \quad (1)$$

The absence of humboldtine (FeC₂O₄·2H₂O) peaks in the XRD spectra of the treated LTS and IWS soils (Figs. 1 and 2) confirms that humboldtine predominantly dissociated following incorporation into the soil matrix. The rapid oxidation of Fe<sup>2+</sup> to Fe<sup>3+</sup> under the prevailing pH and redox conditions drives FeOx dissolution, despite its relatively low solubility (Baran, 2016; Pliego et al., 2014). To verify that oxidative conditions were maintained, redox potential (Eh) measurements were systematically conducted across all leaching tests (Table S4), with Eh values around 400 mV at neutral pH and approximately 200 mV at alkaline pH.

Subsequently, Fe<sup>3+</sup> undergoes hydrolysis to form hydrous (or hydrated) ferric oxides (HFOs), such as ferrihydrite, which precipitate with varying degrees of crystallinity (Brandely et al., 2022b; Tian et al., 2020) according to:

$$Fe^{3+} + 1.75 H_{2}O \rightarrow 0.5 Fe_{2}O_{3} \cdot 0.5H_{2}O(s) + 3 H^{+} \quad (2)$$

in which the release of H<sup>+</sup> leads to acidification. Fe₂O₃·0.5H₂O(s) was selected as a representative average composition of ferrihydrite to account for HFO precipitation in the soils.

No direct evidence of HFOs is observed in the XRD spectra, likely due to their amorphous nature. However, a diffuse background near some diffraction angle of ferrihydrite (53.3°) suggests the presence of such amorphous phases in LTS and IWS after treatment (Figs. 1 and 2). Additionally, the thermogravimetric analysis (TGA) of LTS reveals significant mass loss between 150 and 250 °C, possibly indicating ferrihydrite water loss (circled area 2 in Fig. 3) (Mitov et al., 2002; Subrt et al., 1992). The TGA for IWS, however, is inconclusive due to signal interference in this complex matrix (Fig. S7). Key information on the solid phases can be gained by determining the specific surface by BET (Bi et al., 2025). BET results showed an increase in specific surface area from 12 m²/g to 15.1 m²/g for the LTS treated with 3 % of FeOx, and from 4 m²/g to 9.7 m²/g for the IWS treated with 6 % of FeOx. Despite the relatively low amendment rates, this substantial increase in surface area strongly suggests the formation of a compound with a very high specific surface area (>100 m²/g), consistent with the presence of newly formed and well-dispersed HFOs (Bi et al., 2025). Finally, the color change in both soils to orange/light brown after treatment (Fig. 4) further supports the formation of HFOs, which are known for these characteristic colors (Scheinost and Schwertmann, 1999).

##### 3.1.2. Calcium oxalate precipitation

Calcium oxalate likely forms due to the acidity generated by ferrihydrite precipitation (Eq. (2)), which drives partial calcite dissolution and the release of Ca<sup>2+</sup> ions in the two studied soils (Table 1):

$$CaCO_{3}(s) + 2H^{+} + C_{2}O_{4}^{2-} \rightarrow CaC_{2}O_{4} \cdot 2H_{2}O(s) + CO_{2} \quad (3)$$

Calcium oxalate monohydrate (whewellite) is detected in the XRD spectra of the treated LTS with significant peaks at 14.9° and 24.4° (Fig. 1). Due to background interference in IWS, whewellite presence
<figure><img src="image_3.png" /><figcaption>**Fig. 1.** XRD patterns of the limestone tunnel sludge (LTS) untreated and treated with 8 % FeC₂O₄·2H₂O (FeOx) after 28 days of stabilization time. The color bars reflect the diffraction pics of the key oxalate and iron minerals. (For interpretation of the references to color in this figure legend, the reader is referred to the Web version of this article.)</figcaption></figure>

<figure><img src="image_4.png" /><figcaption>**Fig. 2.** XRD patterns of the industrial wasteland soil (IWS) untreated and treated with 8 % FeC₂O₄·2H₂O (FeOx) after 28 days of stabilization time. The color bars reflect the diffraction pics of the key oxalate and iron minerals. (For interpretation of the references to color in this figure legend, the reader is referred to the Web version of this article.)</figcaption></figure>

cannot be affirmed but a peak at 14.9° seems to corroborate its precipitation (Fig. 2). Notably, other Ca or Mg oxalates are not observed in the spectrum (Figs. 1 and 2). The TGA spectrum of LTS (circled areas 1 and 3 in Fig. 3) further supports whewellite formation, with mass loss observed at 100–150 °C likely due to water loss from the whewellite structure, and a second peak between 450 and 500 °C likely indicating calcium oxalate degradation (Hourlier, 2019).

### 3.1.3. *Quantitative mineralogical change*

The initial mineralogical composition of the untreated soils was determined using a mass balance approach, integrating qualitative phase identification from XRD with quantitative elemental data from XRF and carbon analysis (Table 4). The change in the mineralogy of LTS after the addition of FeOx at 2 and 8 wt% was quantitatively assessed using geochemical modeling (Table 4). The model predicts the complete
<figure><img src="image_5.png" /><figcaption>**Fig. 3.** Derivative weight curve in %/s from TGA analysis of the limestone tunnel sludge (LTS) untreated and treated with 8 % FeC₂O₄·2H₂O (FeOx) after 28 days of stabilization time. The circled and numbered areas highlight regions of interest, which are explained in the text.</figcaption></figure>

dissolution of humboldtine and formation of HFOs and whewellite at both dosages, consistent with the following combined reaction:

$$ \mathrm{FeC_2O_4 \cdot 2H_2O(s) + CaCO_3(s) + 0.25 O_2 \rightarrow CaC_2O_4 \cdot H_2O(s) + 0.5Fe_2O_3 \cdot 0.5H_2O(s) + CO_2 + 0.75H_2O} \quad (4) $$

The calculated contents of ferrihydrite and calcium oxalate are 1.2 wt% and 1.7 wt%, respectively, after addition of 2 wt% FeOx, and 4.6 wt % and 6.3 wt% for 8 wt% FeOx (Table 4). Among the initial minerals, calcite is the only phase that significantly dissolves, whereas dolomite,

the primary Mg source, remains undissolved. Additionally, magnesium oxalate remains undersaturated when CaC₂O₄, CaCO₃ and dolomite coexist.

It is important to emphasize that, in soils lacking pH-buffering phases, such as Ca or Mg carbonates or clays, the addition of FeOx may result in a significant drop in pH. Furthermore, the effective precipitation of Ca or Mg oxalates relies on the presence of readily available Ca or Mg sources (e.g., Ca or Mg carbonates, dissolved in pore solutions). These constraints highlight the need for careful consideration of soil composition when applying FeOx-based treatments, particularly in soils with low buffering capacity, such as sandy soils. Further research is needed to validate this treatment for other soil types. In cases where soil composition is uncertain or lacks buffering phases and/or readily available Ca/Mg sources, a practical approach would be the co-addition

<caption>**Table 4**</caption>

Estimation of the mineral phase assemblage based on XRF and XRD results for the untreated limestone tunnel sludge (LTS) and modeled predictions after FeC₂O₄·2H₂O (FeOx) addition.

<table><thead><tr><th>Content [wt.% of solid mass]</th><th>Untreated LTS</th><th>+2 % FeOx</th><th>+8 % FeOx</th></tr></thead><tbody><tr><td>Calcite</td><td>33.7</td><td>32.0</td><td>27.4</td></tr><tr><td>Dolomite</td><td>50.4</td><td>49.5</td><td>47.2</td></tr><tr><td>Gypsum</td><td>0.9</td><td>0.9</td><td>0.8</td></tr><tr><td>Quartz</td><td>7.9</td><td>7.9</td><td>7.4</td></tr><tr><td>Celestite</td><td>1.5</td><td>1.5</td><td>1.4</td></tr><tr><td>Gypsum</td><td>0.9</td><td>0.9</td><td>0.8</td></tr><tr><td>Anorthite</td><td>1.0</td><td>1.0</td><td>0.9</td></tr><tr><td>Other minor phases <sup>a</sup></td><td>3.7</td><td>3.6</td><td>3.5</td></tr><tr><td>Amorphous ferrihydrite</td><td>0</td><td>1.2</td><td>4.4</td></tr><tr><td>Whewellite</td><td>0</td><td>1.6</td><td>6.1</td></tr></tbody></table>

<figure><img src="image_6.png" /><figcaption>**Fig. 4.** Optical microscope pictures of the limestone tunnel sludge (LTS, Fig. A and B) and industrial wasteland soil (IWS, Fig. C and D) untreated (left) and treated with FeC₂O₄·2H₂O (right).</figcaption></figure>

<footnote><sup>a</sup> anorthite, celestite, goethite, palygorskite.</footnote>
of powered limestone, a cost-effective and widely available material, or other Ca/Mg carbonates to address these issues.

## 3.2. Immobilization of Sb and Mo by FeOx addition

The standard leaching test EN 12457-2 was used as a benchmark for assessing the efficiency of different FeOx dosages in immobilizing molybdate and antimonate. This section presents an analysis of the total released masses normalized to the initial dry mass of soil (mg/kg DM) for liquid-to-solid ratio (L/S) of 10 L/kg DM, as required by environmental threshold standards.

### 3.2.1. FeOx dosage optimization

Regarding Mo, the European inert threshold for soils is 0.5 mg/kg DM (OJEU, 1999). In LTS, this threshold is reached with only 1 wt% FeOx addition, and levels fall below the limit of quantification (LQ = 0.05 mg/kg DM) at dosages of 6 wt% or more (Fig. 5). In IWS, Mo levels are already below the inert threshold, but a 2 wt% addition further reduces Mo release by half, from 0.38 to 0.16 mg/kg DM (Fig. 6).

For Sb, achieving the European inert limit of 0.06 mg/kg DM requires 5 wt% FeOx or above (Fig. 6). The Sb limit is about ten times lower than the Mo one, necessitating a higher treatment level.

A primary objective of this study was to achieve compliance with European inert standards for Mo and Sb. Although the threshold limits were met with FeOx dosages of 1 wt% for LTS and 5 wt% for IWS, slightly higher dosages were selected to ensure consistent compliance and account for potential variability. Natural soil heterogeneity and analytical measurement uncertainties, reflected by the error bars on the graphs, justify this precaution. Therefore, dosages of 2 wt% for LTS and 6 wt% for IWS were chosen for further evaluation of environmental performance in Sec. 3.3 and 3.4.

### 3.2.2. Sb and Mo immobilization mechanisms

The main retention mechanism is assumed to be sorption of molybdate and antimonate oxyanions onto HFOs since this process is well-established in the literature (Shi et al., 2021).

The high efficiency of the treatment can primarily be attributed to the use of finely powdered iron oxalate dihydrate (Sec. 2.2), which allows uniform distribution throughout the soil during blending. Once distributed, $Fe^{2+}$ dissolves, is oxidized and subsequently precipitates as poorly crystalline HFOs throughout the soil matrix (Sec. 3.1.1). This precipitation in a nearly amorphous form is advantageous, as these

<figure><img src="image_7.png" /><figcaption>**Fig. 6.** Effect of $Fe_2O_4 \cdot 2H_2O$ (FeOx) addition on Mo and Sb leaching after 7 days of stabilization time (EN 12457-2 test) in industrial wasteland soil (IWS). FeOx dosage in % of soil dry mass. Untreated = soil without amendment. Dashed lines represent the European regulation for inert waste. Red continuous lines stand for the quantification limits. (For interpretation of the references to color in this figure legend, the reader is referred to the Web version of this article.)</figcaption></figure>

minerals exhibit higher retention capacities for trace metals compared to crystalline oxides, due to their greater specific surface areas that provide abundant adsorption sites (Shi et al., 2021; Tian et al., 2024). The fact that $Fe^{2+}$ goes first in solution before being oxidized also promotes a uniform distribution of HFOs. Fig. 4 shows such a uniform distribution with a consistent orange to light-brown coloration typical of HFOs.

The assumption of Sb and Mo sorption onto HFOs is supported by Figs. 7 and 8, which show a decrease of Mo and Sb concentration with increasing FeOx addition for both soils. The non-linear relationship between Mo and Sb concentrations and FeOx content arises from the limited concentrations of these oxyanions in solution (maximum Mo concentration = 0.15 mg/L in LTS and 0.019 mg/L in IWS, maximum Sb

<figure><img src="image_8.png" /><figcaption>**Fig. 5.** Effect of $Fe_2O_4 \cdot 2H_2O$ (FeOx) addition on Mo leaching after 7 days of stabilization time (EN 12457-2 test) in limestone tunnel sludge (LTS). FeOx dosage in % of soil dry mass. “Untreated” means soil without amendment. Dashed lines represent the European regulation for inert waste. Red continuous lines stand for the quantification limits. (For interpretation of the references to color in this figure legend, the reader is referred to the Web version of this article.)</figcaption></figure>

<figure><img src="image_9.png" /><figcaption>**Fig. 7.** Mo concentrations in eluates of the EN 12457-2 leaching test with increasing addition of $Fe_2O_4 \cdot 2H_2O$ (FeOx) in wt. % of dry mass (DM) soil in limestone tunnel sludge (LTS) and comparison with geochemical modeling.</figcaption></figure>
<figure><img src="image_10.png" /><figcaption>**Fig. 8.** Mo and Sb concentrations in eluates of the EN 12457-2 leaching test with increasing addition of FeC₂O₄·2H₂O (FeOx) in wt. % of dry mass (DM) soil in industrial wasteland soil (IWS).</figcaption></figure>

concentration = 0.038 mg/L in IWS. As the available molybdate and antimonate decrease, additional HFOs no longer proportionally increase sorption. Ultimately, sorption reaches a threshold beyond which additional sorption sites have minimal effect under the present conditions.

The modeling of LTS with FeOx closely aligns with the experimental results (Fig. 7), further confirming molybdate sorption onto HFO surfaces as the predominant retention mechanism. As noted in Sec 2.5, the

modeling was performed exclusively on LTS, as the complex matrix of IWS and its organic matter (OM) content posed significant challenges to accurate modeling (Table 1 and Table S2).

### 3.2.3. Anionic competition

Results from the EN 12457-2 tests (Sec. 3.2.1) showed that the release of sulfate was notably high for LTS (about 3300 mg/kg) and moderate for IWS (about 700 mg/L). However, Mo and Sb showed limited competition with sulfates for sorption in both soils. This is in agreement with the literature indicating that sulfate generally has minimal impact on the sorption of metalloid oxyanions onto Fe(III) oxides or hydroxides (Wang et al., 2021; Xu et al., 2006). This absence of competition is a critical observation, given the frequent presence of sulfates in waste materials.

Similarly, no competitive effects were observed with oxalates. No significant increase in DOC was detected following FeOx addition in both soils (Fig. S8 and Fig. 10), likely due to the rapid precipitation of calcium oxalate solids. This is a favorable outcome, as oxalate could otherwise compete with Sb and Mo oxyanions for sorption sites (Han et al., 2020). However, as highlighted in Sec. 3.1.3, caution is advised when applying this treatment to soils lacking mobilizable Mg and Ca, as oxalate may remain in solution and act as a possible competing species for oxyanion sorption.

As described in Sec. 2.2 the TCLP leaching test involves the addition of acetic acid, introducing about 6000 mg/L of acetate into the solution (pH = 6.2). Despite the high acetate concentration, Sb and Mo retention remained effective in both soils (Fig. 9-D and Fig. 10-E-F). A slight competition effect of acetate was observed in LTS at this elevated acetate concentration. For instance, at pH 6, the Mo concentration reached 0.04 mg/L in TCLP (Fig. 9), whereas it was below 0.01 mg/L in the absence of

<figure><img src="image_11.png" /><figcaption>**Fig. 9.** Effect of stabilization time on pH and the concentration of Fe, Ca and Mo for the 2 wt% FeC₂O₄·2H₂O (FeOx) treatment in limestone tunnel sludge (LTS) according to the leaching standards EN 12457-2 and US EPA 1311 TCLP. The dashed black line represents the European standard inert waste thresholds for Mo. The red continuous line represents the quantification limit (the results plotted on this line are, in fact, equal to or below this limit). The black solid line represents the Mo release of untreated LTS for EN 12457-2 and TCLP. (For interpretation of the references to color in this figure legend, the reader is referred to the Web version of this article.)</figcaption></figure>
<figure><img src="image_12.png" /><figcaption>**Fig. 10.** Effect of stabilization time on pH and the concentration of Ca, DOC, Fe, Mo and Sb for the 6 wt % FeC₂O₄·2H₂O (FeOx) treatment in industrial wasteland soil (IWS) according to the leaching standards EN 12457-2 and US EPA 1311 TCLP. The dashed black line represents the European standard inert waste thresholds for Mo and Sb. The red continuous line represents the quantification limit (the results plotted on this line are, in fact, equal to or below this limit). The black solid line represents the Mo and Sb release of untreated IWS for EN 12457-2 and TCLP. (For interpretation of the references to color in this figure legend, the reader is referred to the Web version of this article.)</figcaption></figure>

acetate during the EPA 1313 test (Fig. 11-B).

Additionally, we systematically measured other potential competitors, including inorganic oxyanions such as phosphate, nitrate, as well as other metallic and metalloid oxyanions (e.g., As, Cr). Their concentrations in solution are reported in Table S3 for LTS and Table S5 for IWS. As shown, these concentrations were sufficiently low, indicating they did not represent significant competitive effect on molybdate sorption onto HFOs.

### 3.3. Time-dependent leaching behavior

The selected treatments chosen to ensure compliance with European inert thresholds, while meeting industrial criteria for cost and efficiency, were 2 wt% FeOx for LTS and 6 wt% FeOx for IWS (Sec. 3.2.1). In the

following sections, an environmental assessment at these dosages is conducted using leaching tests.

#### 3.3.1. General chemistry

During the EN 12457-2 leaching tests, a slight initial pH decrease was observed in the treated soils from 8.4 to 7.2 in LTS and from 7.9 to 6.0 in IWS, after 6 h of stabilization. For longer stabilization times, the pH in both soils returned to values close to their initial levels before treatment (Fig. 9-A and Fig. 10-A). This moderate pH decrease can be attributed to the buffering effect of calcite (Eq. (4)). In addition, although LTS exhibited a slight initial release of Ca from 100 mg/L to 200 mg/L after 6 h, Ca concentrations remained generally constant in both soils over time (Fig. 9-B and Fig. 10-B). This combined stability of pH and Ca concentrations suggests that calcite dissolution and calcium oxalate
<figure><img src="image_13.png" /><figcaption>**Fig. 11.** Acid neutralization capacity (ANC) and effect of pH on the concentration of Mo, Fe and Ca according to the leaching standard US EPA 1313 on untreated Limestone Tunnel Sludge (LTS) soil and after treatment with 2 % $FeC_2O_4 \cdot 2H_2O$ (FeOx). Natural pH at equilibrium stands for pH with no acid or base addition. The red continuous line represents the quantification limits (the results plotted on this line are, in fact, equal to or below this limit). (For interpretation of the references to color in this figure legend, the reader is referred to the Web version of this article.)</figcaption></figure>

precipitation mainly occur within the first 6 h. This result is further supported by the constant DOC concentrations, which show no evidence of increased oxalate concentration. In the LTS, DOC was below the LQ (Fig. S8), while in the IWS, DOC remained close to its initial value of 20 mg/L with no significant variations over time (Fig. 9-C). Under TCLP conditions, there was no significant difference in pH and Ca concentrations before and after treatment in both soils. Additionally, the DOC content was primarily dictated by the use of dilute acetic acid as eluate (Sec.2.2) and could not be considered relevant for further interpretation.

The oxidation of $Fe^{2+}$ to $Fe^{3+}$ was followed by the formation of insoluble HFOs (Eqs. 1 and 2). These two reactions are expected to occur quickly under the present near-neutral pH and oxidizing conditions (Morgan and Lahav, 2007). Indeed, no iron was detected in solution beyond 7 days for either leaching test (Fig. 9-C and Fig. 10-D). In addition, the Fe concentration was negligible at early stabilization times compared to the total Fe introduced by the FeOx treatment. Under EN 12457-2 conditions (L/S = 10 kg/L), FeOx additions of 2 wt% (LTS) and 6 wt% (IWS) correspond to 620 mg/L and 1860 mg/L, respectively. In comparison, the early Fe concentrations in treated LTS and IWS are 0.24 mg/L and 38 mg/L, respectively. Therefore, both EN 12457-2 and TCLP tests support the hypothesis of rapid FeOx dissolution and subsequent formation of HFOs, consistent with the proposed mechanisms in Sec.3.1.

### 3.3.2. Rapid immobilization of Sb and Mo

The literature indicates that Mo and Sb sorption onto HFOs surfaces occurs within 24 h (Sun and Selim, 2019; Tandy et al., 2017). Indeed, the EN 12457-2 leaching tests showed a rapid and strong retention of Sb and Mo in both soils (Fig. 9-D, Fig. 10-E and Fig. 10-F). In the LTS, Mo concentration decreased from 0.15 mg/L to 0.03 mg/L within 6 h, with a slight increase to 0.05 mg/L at 90 days. The latter is likely due to a gradual pH increase from 7.2 to 7.9, which affected HFO surface charges

and thus Mo retention. Overall, there was a 67 % reduction of Mo concentration, remaining within the European inert limit of 0.05 mg/L. In the IWS, Mo dropped from 0.02 mg/L to below the LQ (0.001 mg/L) after the addition of 6 wt% FeOx, which corresponds to a least 90 % reduction. Sb also decreased significantly, from 0.04 mg/L to 0.006 mg/L (88 % reduction). No increase in Mo and Sb concentrations was observed in IWS over time, despite the slow increase in pH.

In TCLP leaching tests, the pH is lower due to acid acetic addition. This decrease in pH enhances oxyanion sorption on HFOs. In the LTS, Mo concentration decreased from 0.15 mg/L to 0.04 mg/L after treatment (73 % reduction). In the IWS, Sb concentration decreased from 0.062 mg/L to 0.005 mg/L (92 % reduction) and Mo concentration dropped from 0.06 mg/L to below the LQ of 0.001 mg/L (i.e. at least a 97 % reduction).

### 3.4. Leaching sensitivity to pH

In addition to the experimental results shown in Figs. 11 and 12, Fig. 13 presents the modeling results of the effect of pH on several elements and minerals in the LTS soil after the addition of 2 wt% FeOx, using the criteria and parameters defined in Sec. 2.5. The simulations were stopped at pH 11, beyond which cementitious phase formation becomes significant and is outside the scope of this study.

#### 3.4.1. Global chemistry

The addition of FeOx had no significant impact on the acid and base neutralization capacity of the soils (Fig. 11-A and Fig. 12-A). The high dolomite and calcite content of the LTS (Table 4) explains the substantial acid demand required to reach the acidic domain. The dissolution of calcite and dolomite at acidic pH yields a significant increase in Ca concentration (Fig. 11-D and Fig. 12-E). However, at pH values of 5.5
and above, the addition of FeOx reduces the concentration of Ca through
the formation of calcium oxalate (whewellite), as corroborated by
geochemical modeling (Fig. 13-A and B).

For both soils, the concentration of Fe exhibits a sharp increase under acidic conditions (pH < 5) and at highly alkaline pH levels (pH > 12), primarily due to the dissolution of iron oxyhydroxides, whether originally present or introduced through the FeOx treatment (Fig. 11-C and Fig. 12-D). Indeed, even before treatment, IWS exhibits considerable Fe release, reflecting its high Fe content (1.24 %) and substantial organic matter content (Table S2). The Fe release in IWS is strongly correlated with DOC (Fig. 12-F), as Fe tends to associate with OM colloids. In contrast, the release of Fe in the LTS remains relatively constant, except under extreme acidic or alkaline conditions. However, the amount of Fe released in both soils remains significantly lower than the total Fe introduced through the FeOx treatment. With 2 wt% FeOx addition in LTS, the theoretical maximum Fe release is 620 mg/L, whereas the highest measured concentration reaches 421 mg/L. Similarly, for the IWS treated with 6 wt% FeOx, the theoretical release is 1860 mg/L,

compared to a maximum observed concentration of 780 mg/L.

Geochemical modeling of the release of Fe from the treated LTS (Fig. 13-B) confirms that the theoretical maximum Fe release is directly associated with the complete dissolution of ferrihydrite at very acidic pH values (pH < 3) (Fig. 13-A). Thus, the lower experimentally observed Fe concentrations are likely constrained by the dissolution kinetics of ferrihydrite within the 24-h duration of the 1313 leaching test. Finally, it is important to emphasize that the significant acid consumption required to lower the pH below 3 in LTS (Fig. 11-A) suggests that HFOs are unlikely to dissolve under typical oxidant acidic conditions encountered in the environment.

3.4.2. Mo and Sb immobilization

A significant decrease in Mo and Sb retention was observed at pH values above 9 in both soils (Fig. 11-B, Fig. 12-B and Fig. 12-C). This trend supports the hypothesis of Sb and Mo sorption onto HFOs. The literature indicates that Fe(III) oxides and hydroxides have peak adsorption capacities at pH 3–5, with diminished capacities at higher pH

<figure><img src="image_14.png" /><figcaption>**Fig. 12.** Acid neutralization capacity (ANC) and effect of pH on the concentration of Sb, Mo, Fe, Ca and DOC according to the leaching standard US EPA 1313 on untreated Industrial Wasteland Soil (IWS) soil and after treatment with 6 % FeC₂O₄·2H₂O (FeOx). Natural pH at equilibrium stands for pH for no acid or base addition. The red continuous line represents the quantification limits (the results plotted on this line are, in fact, equal to or below this limit). (For interpretation of the references to color in this figure legend, the reader is referred to the Web version of this article.)</figcaption></figure>
<figure><img src="image_15.png" /><figcaption>**Fig. 13.** Modeling of the effect of pH on (A) the solid phase content of ferrihydrite and whewellite, (B) the concentrations of Fe and Ca, (C) the concentration of Mo for the LTS soil after addition of 2 wt% $FeC_2O_4 \cdot 2H_2O$ (FeOx) at L/S = 10 L/kg. The horizontal black dashed line represents the limits of quantification LLOQ (the results plotted on this line are, in fact, equal to or below this limit), and the brown continuous line represents the maximum release of Fe for 2 wt% FeOx. (For interpretation of the references to color in this figure legend, the reader is referred to the Web version of this article.)</figcaption></figure>

and little to no sorption above pH 9 (Shi et al., 2021). This is partly related to the point of zero charge of HFOs ($pH_{pzc}$ between 7.6 and 8.5). Above the point of zero charge ($pH_{pzc}$), the surface acquires a negative charge, which hinders the sorption of anionic species such as $MoO_4^{2-}$. In addition, Sb is likely present in solution as $Sb(OH)_6$, which is reversibly bound to HFOs and highly sensitive to pH variations due to both changes in surface charge and competition with hydroxyl ions at alkaline pH conditions (Peng et al., 2023).

Geochemical modeling of the LTS treated with 2 wt% FeOx further confirms the Mo sorption mechanism, showing maximum sorption between pH 3 and 7 and significant desorption above pH 9 (Fig. 13-C). At very acidic pH values (<3), the simulations also indicate that HFO dissolution would theoretically lead to the release of Mo and Sb (Fig. 13-A and Fig. 13-C). However, no clear increase in Mo or Sb was observed experimentally. This discrepancy could be attributed to the partially delayed dissolution of ferrihydrite combined to a stronger sorption affinity at acidic pH as described in Sec. 3.4.1.

The influence of reducing conditions was not examined in this study in which oxidizing conditions prevail (Table S4). However, reducing environments—particularly those driven by microbial activity—could facilitate HFO dissolution and promote recrystallization into more crystalline phases, such as magnetite (Wang et al., 2022a). This transformation may decrease the adsorption capacity of oxyanions. In addition, redox conditions play a crucial role in oxyanion speciation, especially for Sb, as Sb(III) and Sb(V) exhibit distinct chemical behaviors and affinities (Peng et al., 2023). Therefore, the impact and occurrence of reducing conditions should be carefully assessed depending on the intended disposal or valorization strategy.

Overall, Sb and Mo retention remained effective across a wide pH range, particularly under acidic to neutral conditions, which are the most common in natural oxidizing environments (Slessarev et al., 2016).

### 3.4.3. Metal-oxalate complexation

The presence of oxalate in solution can enhance the release of cations such as $Zn^{2+}$ and $Cu^{2+}$ by forming stable aqueous complexes (Krishnamurty and Harris, 1961). However, in this study, no significant increase in Zn or Cu concentrations was observed (Fig. S9), likely due to the immobilization of oxalate as calcium oxalate.

As discussed in Sec. 3.1.3, caution should be exercised when applying this treatment to soils lacking readily available Mg and Ca, particularly in the presence of co-contaminants such as Zn and Cu. In such conditions, oxalate may remain in solution rather than precipitating as low-solubility Ca or Mg oxalate, making it available to complex with Zn, Cu, or other metals, potentially enhancing their mobility and bioavailability.

## 4. Conclusion

This study demonstrates the potential of iron oxalate dihydrate (FeOx) as an effective amendment for the immobilization of Mo and Sb in contaminated soils. Two distinct soil types were evaluated: a limestone tunnel sludge (LTS), primarily composed of carbonates and affected by geogenic Mo, and an industrial wasteland soil (IWS), contaminated with anthropogenic Mo and Sb, and characterized by high feldspar and organic matter content.

Mechanistic insights were obtained through direct characterization (TGA, XRD, optical microscopy), standardized leaching tests (EN 12457-2, US EPA 1311, and 1313), and geochemical modeling. Under near-neutral and oxidizing conditions, FeOx dissociates into $Fe^{2+}$ and $C_2O_4^{2-}$, with $Fe^{2+}$ rapidly oxidized to $Fe^{3+}$, forming poorly crystallized hydrous ferric oxides (HFOs), likely ferrihydrite. The resulting acidification is mitigated by the dissolution of calcite, which releases $Ca^{2+}$ ions that react with $C_2O_4^{2-}$ to precipitate calcium oxalate monohydrate (whewellite).

Mo and Sb retention occurs primarily via sorption onto newly formed HFOs, with effective immobilization achieved at ≥1 wt% FeOx for LTS
and ≥5 wt% for IWS to meet European inert waste thresholds. Leaching tests confirmed the rapid FeOx transformation into HFOs and whe-wellite within the first 6 h, ensuring sustained retention. Mo and Sb leaching was reduced by over 67 % for Mo in LTS with 2 wt% FeOx addition and up to 88 % for Sb and 90 % for Mo in IWS with 6 wt% FeOx. The immobilization of molybdate and antimonate was stable at pH ≤ 8–9, below the point of zero charge (pHpcz) of HFOs, with no compet-itive sorption observed with sulfates or acetates. Additionally, oxalates did not interfere with retention due to their rapid precipitation as insoluble calcium oxalate.

From an industrial and environmental application perspective, the rapid efficiency of FeOx offers a significant advantage. Oxalate biode-gradability and its ability to form highly insoluble complexes with Ca and Mg make it also suitable for such applications. However, certain limitations of this treatment must be highlighted. As with most treat-ments using iron salts, the present FeOx treatment should not be applied to soils lacking pH-buffering phases, as this may result in strongly acidic conditions. In addition, for FeOx treatment specifically, the absence of readily mobilizable Ca and Mg in the soil may increase oxalate mobility and compromise contaminant immobilization in the case of multi-contaminated soils. To address these limitations, particularly in cases where soil composition is uncertain or lacks buffering capacity and Ca/ Mg sources, incorporating powdered limestone or other Ca/Mg car-bonates could be a viable solution. While this approach was not inves-tigated in the current study, further research is needed to evaluate its effectiveness.

Furthermore, this study highlights the need for further investigations to assess the long-term stability of the FeOx treatment under realistic environmental conditions, particularly under variable redox conditions. Microbial activity could induce reducing conditions, potentially leading to HFO dissolution and changes in oxyanion speciation, especially for Sb. To address these uncertainties, pilot-scale experiments (e.g., lysim-eter tests) and scenario-based models for long-term risk assessment in disposal or reuse contexts are essential to validate the treatment's effi-cacy in diverse environmental settings.

### CRediT authorship contribution statement

**Amandine Dumas:** Writing – review & editing, Writing – original draft, Validation, Methodology, Investigation, Formal analysis, Conceptualization.
**Laurent De Windt:** Writing – review & editing, Writing – original draft, Validation, Supervision, Methodology, Formal analysis, Conceptualization.
**Ivan Serclerat:** Writing – review & editing, Validation, Resources, Methodology, Conceptualization.
**David S. Kosson:** Writing – review & editing, Validation, Resources, Methodology.
**Bruno Huet:** Writing – review & editing, Validation, Methodology.
**Florence Sanchez:** Supervision, Resources, Methodology.
**Nouffou Tapsoba:** Writing – review & editing, Validation, Supervision, Resources, Methodology, Funding acquisition.

### 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

The authors would like to thank the managers and technical/ research staff of the analytical team at Holcim Innovation Center and the Department of Civil and Environmental Engineering at Vanderbilt Uni-versity for their analytical assistance and valuable insights, as well as two anonymous reviewers for their constructive comments.

## Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.apgeochem.2025.106626.

### Data availability

Data will be made available on request.

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