# Molybdenum-contaminated soil stabilization/solidification by mixtures of magnesium oxide and limestone calcined clay cement

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

► To cite this version:

Amandine Dumas, Laurent de Windt, Ivan Serclerat, Florence Sanchez, Bruno Huet, et al. Molybdenum-contaminated soil stabilization/solidification by mixtures of magnesium oxide and limestone calcined clay cement. Journal of Hazardous Materials, 2025, 492, pp.138056. 10.1016/j.jhazmat.2025.138056. hal-05042854

HAL Id: hal-05042854

https://minesparis-psl.hal.science/hal-05042854v1

Submitted on 22 Apr 2025

HAL is a multi-disciplinary open access
archive for the deposit and dissemination of sci-
entific research documents, whether they are pub-
lished or not. The documents may come from
teaching and research institutions in France or
abroad, or from public or private research centers.

L'archive ouverte pluridisciplinaire HAL, est
destinée au dépôt et à la diffusion de documents
scientifiques de niveau recherche, publiés ou non,
émanant des établissements d'enseignement et de
recherche français ou étrangers, des laboratoires
publics ou privés.

<figure><img src="image_1.png" /></figure>

Distributed under a Creative Commons Attribution 4.0 International License
<figure><img src="image_2.png" /></figure>

# Molybdenum-contaminated soil stabilization/solidification by mixtures of magnesium oxide and limestone calcined clay cement

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

## HIGHLIGHTS

* Effective treatment of Mo oxyanion-contaminated materials using low CO₂ amendments.

* Comprehensive environmental assessment based on US and EU leaching standards.

* Molybdate sorption on brucite identified as the primary containment mechanism.

* High Mo retention at alkaline pH across varying curing times, even at low dosages.

## GRAPHICAL ABSTRACT

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

## ARTICLE INFO

**Keywords:**

Brucite

Low carbon cement

Metal oxyanion

Modeling

S/S technique

Sustainability

## ABSTRACT

Large quantities of soil excavated from civil engineering projects are found to contain leachable contaminants, including metallic oxyanions like molybdate. This study presents a novel treatment approach by identifying a mixture of magnesium oxide (MgO) and hydraulic binders for the stabilization and solidification (S/S) of molybdenum-contaminated limestone tunnel sludge. XRD analysis and geochemical modeling revealed the presence of brucite in all samples, suggesting that the immobilization mechanism was likely molybdate sorption onto brucite. The addition of cement enhanced the geotechnical properties and improved sulfate retention through the formation of ettringite. Considering efficiency and a low CO₂ footprint, the S/S formulation consisting of 1 wt% MgO and 3 wt% limestone calcined clay cement (LC3) was selected for this study. Environmental assessments were conducted using leaching test standards, including US EPA 1313 to assess pH sensitivity, US EPA 1311 (TCLP) and EN 12457-2 to evaluate the effects of curing time, at intervals from 6 hours to 90 days. Results showed that molybdenum release remained below 0.05 mg/L (the European inert waste threshold) after just 24 hours of curing, with further improvements over time under alkaline conditions. However, the treatment

<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> Holcim - 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@minesparis.psl.eu (L. De Windt).</footnote>
was ineffective at neutral or acidic pH due to brucite dissolution. The S/S of molybdenum-contaminated sludge using MgO and LC3 proved highly efficient for Mo immobilization and is suitable for industrial applications.

# 1. Introduction

Molybdenum (Mo) is an essential trace element for plants and animals, playing a key role in enzymatic redox reactions and various biochemical processes. However, high levels of Mo(VI) in the environment can pose significant health risks, including anemia, liver dysfunction, and osteoporosis [1]. In soils, Mo typically exists as the molybdate anion ($MoO_4^{2-}$), which is chemically stable and primarily found in minerals such as molybdenite and wulfenite. Anthropogenic activities, including metallurgy and agriculture, have contributed to increasing Mo contamination in the environment [1,2].

The mobility of Mo in soils is largely controlled by its adsorption onto mineral and organic surfaces, yet there remain significant gaps in knowledge regarding its behavior, particularly concerning stabilization and solidification (S/S) treatments [1-4]. Most research has focused on the remediation of cationic heavy metals such as copper (Cu), cadmium (Cd) and lead (Pb), while relatively few studies have examined the behavior of oxyanions [5-11]. Among oxyanions, molybdate has received considerably less attention compared to the oxyanions of arsenic (As), chromium (Cr), and antimony (Sb). This represents a critical gap in understanding, especially given the high mobility of $MoO_4^{2-}$ under varying environmental conditions, which complicates its containment.

The need for effective remediation strategies for Mo-contaminated soils is underscored by large-scale civil engineering projects, such as the Grand Paris Express subway, which generate significant volumes of excavated soil with high Mo leachability [12]. Indeed, current regulations, specifically in Europe, categorize these soils based on their leachable contaminant concentrations [13-17]. This presents a significant challenge for waste management and highlights the need for innovative solutions to immobilize Mo and reduce its leachability in compliance with environmental standards.

Magnesium oxide (MgO) has gained attention for its low $CO_2$ footprint and favorable interactions with heavy metals. It has shown high stabilization efficiency for cationic metals such as Cu, Zn and Cd by maintaining an optimal pH range of 9.0-10.5, which reduces the solubility of metal (hydr)oxides [18-20]. Regarding oxyanions, studies have highlighted the role of magnesium hydroxide ($Mg(OH)_2$) as a sorbent for the immobilization of Sb, As, and Cr in various contexts [21-24]. Additionally, As and Cr oxyanions have been specifically studied for their sorption capacity on $Mg(OH)_2$, demonstrating significantly higher affinities compared to non-metallic oxyanions like nitrate ($NO_3^-$) and sulfate ($SO_4^{2-}$) [24]. However, understanding the immobilization mechanisms of Mo using MgO, either alone or in combination with other binders, remains an open area of investigation [25-27].

In addition to MgO, cement binders can be used for immobilizing heavy metals through mechanisms such as precipitation, redox reactions, and incorporation into hydration products [4,8,9]. In the case of molybdate, it can precipitate as powellite ($CaMoO_4$), which controls Mo solubility, although this occurs only at high Mo concentrations [28]. At lower concentrations, more likely in geogenically contaminated soils, molybdate can be adsorbed onto calcium silicate hydrates (C-S-H) or incorporated into mineral phases like ettringite. The similar tetrahedral structures of sulfate and molybdate facilitate this substitution [29-31]. Common binders used in S/S techniques usually include Ordinary Portland Cement (OPC) based binders. However, over the past decade, attention has shifted towards using supplementary cementitious materials (SCMs) such as silica fume, fly ash, and calcined clays to reduce the environmental footprint of cement [9,32]. In particular, limestone calcined clay cement (LC3) has been developed, replacing up to 50 % of clinker with limestone and calcined clay [33,34]. Few studies have been

conducted on the use of LC3 for S/S of solid waste or contaminated soil. A series of articles by Reddy and al. on the use of LC3 for the S/S of Pb-Zn smelter residues and Zn contaminated soil have demonstrated a high immobilization efficiency of LC3 for Zn, Pb and Cd (>88 %) as well as an improvement of the mechanical properties [35-37]. These findings on Zn immobilization efficiency have been corroborated by Wu et al. [38]. Regarding oxyanions, Gu et al. have shown significant retention of Cr following LC3 addition, although the retention mechanisms were not detailed [39]. However, to the best of our knowledge, no studies have been published on Mo retention using LC3, either alone or in combination with other chemical agents.

This paper addresses this knowledge gap by investigating the stabilization and solidification (S/S) of molybdenum in naturally excavated material using a combination of MgO and various cements, including OPC, slag cement (SC), and LC3, representing a novel contribution to the field. The main goals are to identify a suitable mixture of MgO and cement for S/S of molybdate, in line with European standards, and to understand the mechanisms behind Mo immobilization. XRD analysis, saturation index calculations, and mineral modeling were used to identify the hydration products formed with each mixture. A thorough evaluation was performed to assess chemical durability and stabilization effectiveness, using standard leaching tests. US EPA 1313 was applied to evaluate pH sensitivity (range 2-13), while US EPA 1311 (TCLP) and EN 12457-2 were used to measure stabilization over time, with curing durations ranging from 6 hours to 90 days, to ensure both quick efficiency and long-term stability.

The findings provide valuable insights into Mo retention mechanisms, essential for improving remediation strategies and supporting the reuse of contaminated soils in construction, while confirming the formulation's suitability for industrial applications within a realistic pH range and a short curing period.

# 2. Materials and methods

## 2.1. Contaminated soil and treatment additives

### 2.1.1. Mo-contaminated soil

The soil used in this experiment (Fig. 1) was sourced from a limestone tunnel sludge generated during the subway extension in Paris, France (Grand Paris Express Project). Based on XRD and XRF analysis, this material is predominantly composed of carbonates dolomite and calcite (Table 1). Gypsum was also detected, and even at low concentrations, it can be a source of sulfate release due to its relatively high solubility.

Following extraction by a tunnel-boring machine, the material underwent several processing steps commonly applied to tunnel fines. Initially, lime was added to reduce its water content and facilitate handling. The material was then stored in a pile for an extended period before being air-dried. Finally, prior to analysis in this study, the material was ground (<4 mm) and homogenized. Given the long exposure to air of the sludge, oxidizing conditions are presumed.

Table 2 gives the elemental bulk contents in major oxides and trace elements of the untreated natural soil. The total molybdenum is present in trace amounts, measured at 7 mg/kg of dry matter (DM). Molybdate is the most likely Mo species under oxidizing conditions [1,2].

### 2.1.2. Reagents

The magnesium oxide (MgO, reagent grade, >99.7 %) was supplied by Sigma Aldrich and consisted of periclase.

The cementitious materials used in this study included ordinary white Portland cement (OPC, CEM I 52.5 N White, Le Teil, France) and
slag cement (SC, CEM III/B 42.5 N, La Malle, France), both supplied by Holcim Company. The SC comprises 72 wt% of blast furnace slag. Hydrated lime ($Ca(OH)_2$) was sourced from St Hilaire Company.

The limestone calcined clay cement (LC3) was specifically prepared by blending 50 wt% OPC (CEM I 52.5 R, Holcim, Saint Pierre la Cour, France), 33 wt% calcined clay (TechZero, Holcim, Saint Pierre la Cour, France), and 17 wt% calcite filler (BL 200 Orgon, Omya, France). Notably, the calcined clay used in LC3 contains 36 wt% amorphous metakaolin.

These cementitious materials were selected for their specific characteristics: OPC as a widely used reference material, SC as a cement commonly employed in S/S treatments of soils, and LC3 as a promising cement due to its low $CO_2$ footprint and enhanced reactivity. Detailed anhydrous and oxide compositions of the cementitious materials are presented in Table 3 and Table 4.

## 2.2. Stabilization/solidification protocol

### 2.2.1. Treatment overview

The first objective was to select the most suitable treatment for Mo retention. The specific reagent quantities and combinations tested are detailed in Table 5. Initially, various percentages of MgO from 0.5 wt% to 5 wt% were evaluated to assess its Mo immobilization capacity. Subsequently, different hydraulic binders were tested at a dosage of 5 % of soil dry mass (DM). This amount was chosen as benchmark as it provides a balance between achieving sufficient reagent content for X-ray diffraction (XRD) mineral phase analysis while maintaining industrial feasibility (cost). Following this, several MgO/LC3 mixtures were examined to optimize the formulation. The S/S protocol involved mechanically mixing contaminated soil with the selected reagents for 3 minutes. The water content was adjusted to achieve the optimum moisture content, determined according to the standard Normal Proctor compaction test [40].

A curing period of 7 days was selected as a compromise to allow sufficient time for Mo immobilization while accommodating the need for rapid industrial application [41]. During this curing phase, samples were stored at 22 ± 1°C in sealed plastic bags to prevent exposure to air and moisture loss. After curing, the treatment effectiveness was evaluated by comparing results with the European regulation for inert wastes (0.5 mg/kg DM) [14], using the standardized batch leaching test EN 12457-2 [15]. Consequently, the figures show the total released normalized to the initial DM of the soil (mg/kg DM), considering the liquid-to-solid ratio (L/S) applied in the associate leaching test.

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

Mineralogical composition of the excavated soil based on XRF and XRD results.</caption><thead><tr><td>Minerals</td><td>Formula</td><td>Bulk content (wt%)</td></tr></thead><tbody><tr><td>Dolomite</td><td>CaMg(CO<sub>3</sub>)<sub>2</sub></td><td>51</td></tr><tr><td>Calcite</td><td>CaCO<sub>3</sub></td><td>34</td></tr><tr><td>Quartz</td><td>SiO<sub>2</sub></td><td>8</td></tr><tr><td>Gypsum</td><td>CaSO<sub>4</sub>·2 H<sub>2</sub>O</td><td>&lt; 1</td></tr><tr><td>Celestite</td><td>SrSO<sub>4</sub></td><td>&lt; 1</td></tr><tr><td>Falcondoite (sepiolite)</td><td>(Ni,Mg)<sub>4</sub>Si<sub>6</sub>O<sub>15</sub>(OH)<sub>2</sub>·6 H<sub>2</sub>O</td><td>&lt; 1</td></tr><tr><td>Anorthite (feldspar)</td><td>Al<sub>2</sub>Ca(SiO<sub>4</sub>)<sub>2</sub>/Al<sub>2</sub>O<sub>3</sub>·CaO·2SiO<sub>2</sub></td><td>&lt; 1</td></tr><tr><td>Palygorskite (clay)</td><td>Mg<sub>5</sub>(Si,Al)<sub>8</sub>O<sub>26</sub>(OH)<sub>2</sub>·8 H<sub>2</sub>O</td><td>&lt; 1</td></tr><tr><td>Hematite/Goethite</td><td>Fe<sub>2</sub>O<sub>3</sub> / FeO(OH)</td><td>&lt; 1</td></tr></tbody></table>

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

Mean bulk composition of major and trace elements in the untreated soil (dry mass, DM). $LOI = Loss of Ignition$.</caption><thead><tr><td>Major oxide and Carbon</td><td>Content (wt %)</td><td>Trace element</td><td>Content (mg/kg of DM)</td></tr></thead><tbody><tr><td>LOI</td><td>40.0</td><td>As</td><td>&lt; 1</td></tr><tr><td>SiO<sub>2</sub></td><td>9.9</td><td>Ba</td><td>111</td></tr><tr><td>Al<sub>2</sub>O<sub>3</sub></td><td>1.4</td><td>Be</td><td>&lt; 1</td></tr><tr><td>Fe<sub>2</sub>O<sub>3</sub></td><td>0.7</td><td>Cd</td><td>1.7</td></tr><tr><td>CaO</td><td>33.7</td><td>Co</td><td>&lt; 1</td></tr><tr><td>MgO</td><td>10.5</td><td>Cr</td><td>20.5</td></tr><tr><td>K<sub>2</sub>O</td><td>0.3</td><td>Cu</td><td>8.3</td></tr><tr><td>Na<sub>2</sub>O</td><td>0.2</td><td>Hg</td><td>&lt; 2</td></tr><tr><td>SO<sub>3</sub></td><td>2.1</td><td>Li</td><td>24</td></tr><tr><td>TiO<sub>2</sub></td><td>0.1</td><td>Mn</td><td>103</td></tr><tr><td>Mn<sub>2</sub>O<sub>3</sub></td><td>&lt; 0.05</td><td>Mo</td><td>6.7</td></tr><tr><td>P<sub>2</sub>O<sub>5</sub></td><td>&lt; 0.05</td><td>Ni</td><td>7.8</td></tr><tr><td>Cr<sub>2</sub>O<sub>3</sub></td><td>&lt; 0.05</td><td>Pb</td><td>&lt; 1</td></tr><tr><td>ZrO<sub>2</sub></td><td>&lt; 0.05</td><td>Sb</td><td>&lt; 4</td></tr><tr><td>SrO</td><td>0.8</td><td>Se</td><td>&lt; 1</td></tr><tr><td>Organic carbon</td><td>0.5</td><td>Sn</td><td>&lt; 1</td></tr><tr><td>Inorganic carbon</td><td>10.4</td><td>Sr</td><td>6390</td></tr><tr><td></td><td></td><td>Te</td><td>&lt; 1</td></tr><tr><td></td><td></td><td>Ti</td><td>171</td></tr><tr><td></td><td></td><td>Tl</td><td>&lt; 1</td></tr><tr><td></td><td></td><td>V</td><td>28</td></tr><tr><td></td><td></td><td>Zn</td><td>38</td></tr></tbody></table>

### 2.2.2. Leaching assessment as a function of curing time and pH

The second objective was to quantitatively evaluate the effect of hydration kinetics and pH changes on Mo retention capacity. Similar to the previous protocol, the treatment was applied by mechanically mixing the soil and reagents for 3 minutes. Samples were then collected

<figure><img src="image_4.png" /><figcaption>**Fig. 1.** Photographs of the untreated soil in full size view (cm scale, left) and optical microscope view (µm scale, right).</figcaption></figure>
<table><caption>**Table 3**
Mineralogical composition of the cementitious reagents (XRD).</caption><thead><tr><th>Bulk content (wt%)</th><th>OPC (white)</th><th>SC</th><th>LC3</th></tr></thead><tbody><tr><td>Amorphous</td><td>-</td><td>61.1</td><td>13.8</td></tr><tr><td>C3S</td><td>67.3</td><td>19.5</td><td>32.3</td></tr><tr><td>C2S</td><td>21.9</td><td>4.9</td><td>5.9</td></tr><tr><td>C3A</td><td>2.4</td><td>2.6</td><td>2.4</td></tr><tr><td>C4AF</td><td>0.6</td><td>2.2</td><td>5.1</td></tr><tr><td>Portlandite</td><td>1.2</td><td>0.6</td><td>-</td></tr><tr><td>Quartz</td><td>-</td><td>-</td><td>17.2</td></tr><tr><td>Gypsum</td><td>-</td><td>-</td><td>0.7</td></tr><tr><td>Semi-hydrate</td><td>-</td><td>1.1</td><td>1.1</td></tr><tr><td>Anhydrite</td><td>1.8</td><td>3.5</td><td>-</td></tr><tr><td>Calcite</td><td>2.9</td><td>-</td><td>17.9</td></tr></tbody></table>

<table><caption>**Table 4**
Oxide composition of the reagents.</caption><thead><tr><th>Oxide composition (wt%)</th><th>MgO</th><th>Ca(OH)<sub>2</sub></th><th>OPC</th><th>SC</th><th>LC3</th></tr></thead><tbody><tr><td>LOI</td><td>/</td><td>24.4</td><td>2.3</td><td>0.6</td><td>8.7</td></tr><tr><td>SiO<sub>2</sub></td><td>/</td><td>0.4</td><td>23.3</td><td>30.3</td><td>35.8</td></tr><tr><td>Al<sub>2</sub>O<sub>3</sub></td><td>/</td><td>0.1</td><td>2.6</td><td>8.1</td><td>8.4</td></tr><tr><td>Fe<sub>2</sub>O<sub>3</sub></td><td>/</td><td>0.4</td><td>0.2</td><td>1.0</td><td>1.8</td></tr><tr><td>CaO</td><td>/</td><td>74.0</td><td>67.3</td><td>49.2</td><td>41.4</td></tr><tr><td>MgO</td><td>> 99.7</td><td>0.5</td><td>0.5</td><td>4.8</td><td>0.5</td></tr><tr><td>K<sub>2</sub>O</td><td>/</td><td>/</td><td>0.4</td><td>0.3</td><td>0.8</td></tr><tr><td>Na<sub>2</sub>O</td><td>/</td><td>/</td><td>0.5</td><td>0.5</td><td>0.1</td></tr><tr><td>SO<sub>3</sub></td><td>/</td><td>0.1</td><td>2.7</td><td>4.6</td><td>1.8</td></tr><tr><td>TiO<sub>2</sub></td><td>/</td><td>/</td><td>0.1</td><td>0.4</td><td>0.2</td></tr><tr><td>Mn<sub>2</sub>O<sub>3</sub></td><td>/</td><td>/</td><td>/</td><td>0.2</td><td>0.1</td></tr><tr><td>P<sub>2</sub>O<sub>5</sub></td><td>/</td><td>/</td><td>0.1</td><td>0.1</td><td>0.2</td></tr><tr><td>Cr<sub>2</sub>O<sub>3</sub></td><td>/</td><td>/</td><td>/</td><td>/</td><td>/</td></tr><tr><td>ZrO<sub>2</sub></td><td>/</td><td>/</td><td>/</td><td>/</td><td>/</td></tr><tr><td>SrO</td><td>/</td><td>0.1</td><td>0.1</td><td>0.1</td><td>/</td></tr></tbody></table>

<table><caption>**Table 5**
Summary of the studied treatments (stabilizing agent and dosage) in percentage of soil dry mass (DM).</caption><thead><tr><th>Treatment code</th><th>MgO (wt%)</th><th>OPC (wt%)</th><th>SC (wt%)</th><th>LC3 (wt%)</th><th>Ca(OH)<sub>2</sub> (wt%)</th></tr></thead><tbody><tr><td>MgO</td><td>0.5, 0.8, 1, 1.5, 3, 5</td><td>-</td><td>-</td><td>-</td><td>5</td></tr><tr><td>Ca(OH)<sub>2</sub></td><td>-</td><td>-</td><td>-</td><td>-</td><td>-</td></tr><tr><td>OPC</td><td>-</td><td>5</td><td>-</td><td>-</td><td>-</td></tr><tr><td>SC</td><td>-</td><td>-</td><td>5</td><td>-</td><td>-</td></tr><tr><td>LC3</td><td>-</td><td>-</td><td>-</td><td>5</td><td>-</td></tr><tr><td>MgO + OPC</td><td>5</td><td>5</td><td>-</td><td>-</td><td>-</td></tr><tr><td>MgO + Ca(OH)<sub>2</sub></td><td>5</td><td>-</td><td>-</td><td>-</td><td>5</td></tr><tr><td>MgO + SC</td><td>5</td><td>-</td><td>5</td><td>-</td><td>-</td></tr><tr><td>MgO + LC3</td><td>1<br>3<br>5</td><td>-<br>-<br>-</td><td>-<br>-<br>-</td><td>1, 3, 5<br>1, 3, 5<br>5</td><td>-<br>-<br>-</td></tr></tbody></table>

after 6 h, 24 h, 48 h, 7d, 14d, 28d, 70d and 90 days of curing time to evaluate treatment kinetics using the EU and US leaching test EN 12457-2 and US EPA 1311 (TCLP) [16]. To assess pH sensitivity, the US EPA 1313 leaching test [42] was used after 28 days of curing. During curing, the samples were stored at 22 ± 1 °C in zip-lock plastic bags enclosed within an aluminum coated bag to prevent any interaction with air (in particular to avoid carbonation) and water loss over the long term. Table 6 summarizes the characteristics of the leaching tests and the corresponding curing. All leaching tests used are standardized batch tests; additional details on the protocols can be found in the referenced sources. The same leaching tests were also conducted on the untreated soil for comparison.

All figures in Section 3.2 display sulfate release. Elemental sulfur (S), measured by ICP-OES, was converted to sulfate due to the oxidizing conditions (see Section 2.1.1) and the presence of sulfate-bearing minerals such as celestite and gypsum in the studied material. This

<table><caption>**Table 6**
Leaching test characteristics and curing times for leaching assessment.</caption><thead><tr><th>Leaching test</th><th>Characteristics</th><th>Curing time (day)</th></tr></thead><tbody><tr><td>EN 12457-2 [15]</td><td>European compliance test, granular materials in batch, 24 h contact time, Liquid-to-Solid ratio (L/S) = 10 L/kg, natural pH equilibrium</td><td>0.25, 1, 2, 7, 28, 70 and 90</td></tr><tr><td>1311 [16]</td><td>Compliance test US EPA standard TCLP, granular materials in batch, L/S= 20 L/kg, 18 h contact time, acetic acid addition</td><td>1, 7 and 28</td></tr><tr><td>1313 [42]</td><td>US EPA characterization method, granular materials in batch, L/S= 10 L/kg, 24 h contact time with 9 imposed final extract pH values from 2 to 13</td><td>28</td></tr></tbody></table>

conversion facilitates direct comparison with Section 3.1, where sulfate was measured by ion chromatography.

## 2.3. Geotechnical testing

To monitor the geotechnical properties and assess the strength and durability of the treated soils, the California bearing ratio (CBR) and unconfined compressive strength (UCS) tests were conducted according to NF EN 13286-47 [43] and NF EN 13286-41 [44], respectively. The CBR is a penetration test used to evaluate the mechanical strength by measuring the immediate bearing index (IBI), obtained immediately after sample preparation, and the CBR immersion index (iCBR), measured after 4 days of immersion in water. The UCS test was used to study the mechanical performance of both untreated and stabilized contaminated soil samples after a curing period of 7 and 60 days at 20 °C and 90 % relative humidity.

## 2.4. Analytical methods

### 2.4.1. Solid phases

The elemental composition of the untreated soil was determined using X-ray fluorescence (XRF) with wavelength dispersion sequential analysis (S8 TIGER Serie 2 from Brucker) and mineralization by acid attack with *aquia regia* assisted by microwave digestion (ISO 54321:2020) followed by ICP-OES analysis (ISO 11885:2009) using Avio 500DV from PerkinElmer. The total organic and inorganic carbon contents (TOC and TIC) were analyzed by using a combination of combustion under nitrogen (120°C) and oxygen (1100°C) and infra-red (IR) detection on the C/H analyzer RC612 from LECO. The mineral composition of powdered samples (<60 µm) was determined using qualitative X-Rays diffraction (XRD) using D8 Advance from Brucker (copper tube, power: 40 mA × 25 kV, continuous scan from 3 to 65°, program 10 min, steps: 4336).

### 2.4.2. Eluate

The filtered eluate (0.45 µm, PVDF) underwent analysis for the content of metals and metalloids using ICP-MS/OES following the EN ISO 11885:2009 methodology on an Avio 500DV instrument from PerkinElmer. Additionally, anions such as sulfate were monitored using ion chromatography in accordance with ISO 15923-1:2013 standards, utilizing HPIC Thermo Scientific Dionex ICS-6000. pH measurements were conducted using SenTix 41 probe from WTW.

## 2.5. Geochemical modeling

The geochemical module CHESS of the HYTEC code [45] was used to calculate the saturation indices of the batch solutions as well as the mineralogical assemblies for the different recipes at 25°C. The thermodynamic data came from the Thermoddem V1.10 (2020) database [46]. The stoichiometric formula of metakaolin was added assuming a fully dehydrated form as Al₂Si₂O₇. The B-dot model was employed to correct
for activity at ionic strengths up to 1 molal.

The saturation index (SI) of a given mineral is defined as:

$$SI = \log\left(\frac{IAP}{K_s}\right) \qquad (1)$$

where IAP is the ion activity product and Ks is the solubility product thermodynamic constant. SI values are good indicators for estimating the saturation state of complex aqueous solution such as the present leaching tests versus mineral reactivity. SI values between -0.5 and 0.5 mean that the solid phase is close to thermodynamic equilibrium with the solution. A positive SI indicates that the solution is oversaturated and that the solid phase can precipitate from the solution. Negative SI means that the solid phase cannot precipitate but can possibly dissolve if in contact with the solution.

The mineralogical assemblies of the untreated soil were estimated based on semi-quantitative data from XRD and quantitative analysis from XRF (Table 1). The calculated saturation indices were also taken into account. All solid reactive phases were considered to be at thermodynamic equilibrium. Due to the short curing time (7 days) and contact time in batch (24 hours), the siliceous materials of the soil (quartz, anorthite and palygorskite) were considered as non-reactive since their dissolution are very slow [47]. The dissolution of dolomite was also set as unreactive in the presence of LC3, as discussed in Section 4.2. The treatment additives (MgO, clinker oxides, and metakaolin) were assumed to have fully reacted. A sensitivity analysis was done assuming that only 75 % of the clinker and 50 % of the metakaolin had reacted after 7 days, according to the literature [48,49].

# 3. Results

## 3.1. Treatment selection

The standard leaching test EN 12457-2 [15] was used as a benchmark to evaluate the effectiveness of different treatments in stabilizing molybdate. Sulfate release was also assessed, as it shares a similar oxyanion structure with Mo, making it a potential competing anion. Additionally, sulfate is a regulated contaminant with established environmental thresholds. With respect to other chemical competitors, the release of trace contaminants. (Table 2) from the sludge, including metal cations (e.g. Cu, Ni, Zn) and oxyanions (e.g. Cr), was consistently very low (Table S1). These low releases could not have any significant interactions affecting the S/S response of the amendments in relation to Mo retention.

### 3.1.1. Effect of MgO addition

The incorporation of MgO significantly reduced Mo leaching to levels below the European Union (EU) inert regulated value of 0.5 mg/kg DM (Fig. 2). Fig. 3 further illustrates Mo concentration in eluate for various additions of MgO (wt% of Dry Mass (DM) of the soil). Increasing the MgO proportion lowered the concentration of leached Mo. Even the lowest dosage of 0.5 wt% MgO was effective in reducing the leached concentration of Mo from 1.5 mg/kg DM in the untreated soil to below the EU inert waste threshold. In contrast, the leached sulfate concentration did not fall below the EU inert waste threshold of 1000 mg/kg DM, despite a slight reduction compared to the untreated soil, regardless of the amount of MgO added.

Fig. 2 also shows that the hydration of periclase increased the pH of the batch solution from 8.4 to 10.4 after 7 days. This rise in pH was consistent with the X-ray diffraction spectrum, which showed the disappearance of the periclase phase (MgO) and the formation of brucite (Mg(OH)₂) (Fig. 4). The spectra from 7 days and 90 days of curing exhibited remarkable similarity, indicating that the mineralogical assemblage stabilized within just 7 days. Overall, all MgO quickly hydrated to form brucite, which remained stable over time [50,51] and maintained an equilibrium with the solution, buffering the pH at 10.4.

<figure><img src="image_5.png" /><figcaption>Fig. 2. Effect of MgO addition on sulfate (top) molybdenum (bottom) leaching after 7 days (EN 12457-2 test). MgO dosage in wt% of dry soil (DM). Dashed lines represent the European regulation for inert wastes. Red continuous lines indicate the quantification limits.</figcaption></figure>

<figure><img src="image_6.png" /><figcaption>Fig. 3. Mo concentration in eluate of the EN 12457-2 leaching test for various addition of MgO (wt% of Dry Mass (DM) of the soil).</figcaption></figure>

### 3.1.2. Effect of cementitious materials and MgO addition

The addition of cementitious materials to MgO was evaluated to further enhance Mo stabilization, as well as to improve solidification and geotechnical properties. Eventually, cement may also decrease the leaching of sulfate.

Fig. 5 clearly indicates that no cement binders were able to decrease Mo leaching when used alone. Mixing the contaminated soil with cementitious materials alone showed no significant improvement in Mo retention, and in most cases, had a detrimental effect (Fig. 5). In contrast, the addition of MgO significantly improved Mo stabilization, with the best efficiency observed when mixed with LC3 or SC cements. Regarding sulfate, the addition of cement significantly improved its retention but without decreasing its concentration below the EU inert waste threshold (Fig. 5). The best immobilization of sulfate was obtained with 5 % LC3 alone, bringing the concentration very close to the EU inert waste threshold of 1000 mg/kg DM. Notably, mixing both cements and MgO slightly reduced sulfate retention. In the case of 5 % LC3, the concentrations of sulfate were approximately 1100 mg/kg DM without MgO and 1900 mg/kg DM with MgO.

XRD analysis revealed the presence of brucite in all treatments, even
<figure><img src="image_7.png" /><figcaption>Fig. 4. XRD patterns of the untreated soil and the soil treated with 5 wt% MgO after 7 days and 90 days) of curing.</figcaption></figure>

<figure><img src="image_8.png" /><figcaption>Fig. 5. Comparative effect of hydrated lime (Ca(OH)₂), limestone calcined clay cement (LC3), Portland cement (OPC), and slag cement (SC) on molybdenum and sulfate leaching after 7 days (EN 12457-2 test). Binder dosage in wt% of dry soil. Dashed lines represent the European regulations for inert waste. Red continuous lines indicate quantification limits.</figcaption></figure>

with the addition of cement and the significant increase in pH to around 12 (Fig. 6). Similar to the MgO-only treatment, brucite stabilized within 7 days. Ettringite was detected by XRD after 90 days, but not at 7 days (Fig. 6). However, other common cementitious phases such as portlandite [9] and hydrotalcite, a magnesium hydroxy-aluminate phase commonly formed in cement [52], were not detected by XRD.

### 3.1.3. Effect of LC3 and MgO proportions

The blend of MgO and LC3 cement was selected for efficient retention of Mo (and, to a lesser extent, sulfates), as well as the lowest CO₂ footprint. The effects of different proportions of LC3 and MgO on the untreated soil were then assessed using the standard leaching test EN 12457-2.

Fig. 7 shows that the leaching of Mo was primarily influenced by the quantity of MgO added, rather than by the LC3 content, which is consistent with the results in Section 3.1.2. An increase in the MgO content resulted in a corresponding decrease in the release of Mo. In contrast, the amount of LC3 added had a minimal effect on Mo release but a direct influence on sulfate release (Fig. 7). Increasing the LC3 content enhanced sulfate stabilization, with only the treatment containing 5 % LC3 reaching levels close to the EU threshold for sulfate. Additionally, the pH became more alkaline when the LC3 content increased.

In addition to leaching tests, the geotechnical improvements from adding LC3 to MgO in the formulation of 1 wt% MgO and 3 wt% LC3 were evaluated. These results were compared with those obtained from 1 wt% MgO alone and the untreated soil (Table 7). The ratio of the soaked CBR (iCBR) to the immediate bearing index (IBI) of the untreated soil (about 1) increased by a factor of 4 with the addition of 3 % LC3. The UCS parameter of the untreated soil improved by a factor 5 with MgO and by a factor of 10 with the MgO-LC3 blend.

The “1 % MgO + 3 % LC3” blend provides very efficient chemical stabilization of molybdenum, moderate sulfate retention, and reasonable mechanical performance, all while maintaining a relatively low cost and CO₂ footprint.

## 3.2. Effect of curing time and pH on Mo stabilization

### 3.2.1. Effect of curing time

The duration of curing is usually a key factor in engineering operations due to time constraints. According to the EN 12457-2 standard (EU), the efficiency of Mo immobilization using 1 % MgO and 3 % LC3 improved with increased curing time, though it was already quite effective after just 6 hours of curing (Fig. 8-A). Specifically, after 6 hours, a 67 % reduction in Mo leaching was observed, and by the final measurement at 90 days, this reduction reached 93 %. These results demonstrate the high efficiency of this treatment in both the short and long term. The pH increased dramatically following the addition of the MgO + LC3 treatment, increasing from 8.4 in the untreated soil to 12, then gradually decreased, stabilizing at around 11.2 after 60 days (Fig. 8-B).
<figure><img src="image_9.png" /><figcaption>Fig. 6. XRD patterns of the untreated soil and soil treated with 5 % MgO + 5 % LC3, 5 % MgO + 5 % SC and 5 % MgO + 5 % OPC after 7 and 90 days of curing.</figcaption></figure>

<figure><img src="image_10.png" /><figcaption>Fig. 7. Effect of mix dosage of MgO and limestone calcined clay cement (LC3) on molybdenum and sulfate leaching after 7 days (EN 12457–2 test). Binder dosage in wt% of dry soil. Dashed lines represent the European regulations for inert waste. Red continuous lines indicate quantification limits.</figcaption></figure>

This stabilization trend was further illustrated by the decreased leaching of sulfate over time (Fig. 8–C). Cement hydrates incorporating sulfate like ettringite [8] were progressively formed over time. It should be noted that the initial total concentration of sulfate was higher after treatment addition due to the presence of semi-hydrate and gypsum in LC3 (Table 3). Finally, the Mg concentration was below the lower limit of quantification (LLOQ) at all curing times (Fig. 8–D).

The TCLP standard (US) demonstrated lower Mo retention at all curing times than EN 12457–2 (Fig. 8–A), likely due to the significant eluate pH decrease from 11.6 to 7 following the addition of acetic acid. Nevertheless, an improvement was observed in comparison with the untreated soil, which can be attributed to the buffering effect of the

<caption>Table 7</caption>

California bearing ratio (CBR) and unconfined compressive strength (UCS) test results after 7 days of curing for the untreated soil, the soil treated with 1 % MgO, and soil treated with 1 % MgO + 3 % LC3. IBI = immediate bearing index, iCBR = soaked CBR.

<table><thead><tr><th>Samples</th><th>IBI</th><th>Soaked CBR (iCBR)</th><th>iCBR/IBI ratio</th><th>UCS (MPa) 7 days</th><th>UCS (MPa) 60 days</th></tr></thead><tbody><tr><td>Untreated</td><td>1.6</td><td>1.9</td><td>1.2</td><td>0.06</td><td>0.06</td></tr><tr><td>1 % MgO</td><td>4.3</td><td>10.8</td><td>2.5</td><td>0.29</td><td>0.29</td></tr><tr><td>1 % MgO + 3 % LC3</td><td>7.2</td><td>28.2</td><td>3.9</td><td>0.57</td><td>0.64</td></tr></tbody></table>

treatment; the final eluate pH before treatment was 6.2, compared to 7 after treatment (Fig. 8–B). The higher release of Mg after treatment can be explained by the dissolution of Mg-bearing phases, which are more concentrated due to the addition of MgO (Fig. 8–D).

### 3.2.2. Effect of pH

The sensitivity of Mo retention with respect to the pH range was further assessed using the EPA method 1313 before and after treatment (Fig. 9). The quantity of nitric acid required to lower the eluate pH to less than 6 was significantly high (> 0.9 kgHNO₃/kg DM) (Fig. 9-A). This can be attributed to the mineralogical composition of the studied soil (Table 1) that predominantly consisted of calcite and dolomite. These two carbonate phases, which can be considered as bases from an acid/base chemistry point of view, had a strong buffering capacity [47]. Moreover, the acid addition required was similar before and after treatment, with only a slight increase in acid demand observed at alkaline pH levels (8–11.5), due to the dissolution of the additional cement hydrates.

Fig. 9-B focuses on Mg release across the entire pH range. The Mg concentration was very low at the alkaline pH induced by MgO hydration and cementitious phases. However, when the pH dropped below 8, the concentration of Mg in the eluate significantly exceeded that observed in the untreated soil. This phenomenon is likely attributable to the dissolution of brucite formed following the addition of MgO. This
<figure><img src="image_11.png" /><figcaption>**Fig. 8.** Curing time sensitivity of 1 % MgO + 3 % LC3 on (a) Mo leaching, (b) pH, (c) sulfate leaching, and (d) Mg release according to the leaching standards EN 12457–2 and US EPA 1311 TCLP. Dashed black lines represent the European standard inert waste thresholds for molybdenum. Red continuous lines indicate the quantification limit if applicable.</figcaption></figure>

hypothesis is supported by the saturation indices (SIs) of brucite, as illustrated in Fig. 10. The data indicate clearly negative SIs at pH values below 8, confirming the dissolution of brucite under these conditions. Regarding sulfate, a significant reduction in the total concentration was observed between pH 11 and 12.5 for the treated soil (Fig. 9-C).

Molybdenum retention (Fig. 9-D) was significantly improved at pH levels above 9 in the case of MgO/LC3 addition compared to the untreated soil, in agreement with the previous results. When the pH decreased below 9, the stabilization plateaued, reaching maximum values of approximately 0.2 mg/L for the treated soil and 0.27 mg/L for the untreated soil. Retention increased again between pH 7 and 4 and then decreased again at the very acidic pH values. Given that the studied soil contains a total molybdenum content of 7 mg/kg DM, a complete dissolution in the 1313 leaching test with an L/S ratio of 10 would result in a Mo concentration of 0.7 mg/L. Fig. 9-D, shows a maximum leachable concentration of 0.27 mg/L for the untreated soil, indicating that only 39 % of the molybdenum was mobilized, with the majority remaining stable across the pH range.

# 4. Discussion

## 4.1. Brucite as the main molybdate sorbent

In this study, brucite (Mg(OH)₂) was identified in all samples treated with 5 % MgO, both with and without cement addition, and remained

stable over time, as evidenced by consistent XRD peak intensities between 7 and 90 days of curing (Fig. 4, Fig. 6). This suggests that brucite Mg(OH)₂ plays a key role in molybdenum retention. To the best of our knowledge, there is no study on MoO₄²⁻ sorption on brucite. But these findings align with previous studies that have demonstrated the potential of Mg(OH)₂ for oxyanion sorption [21,23,24,53]. In particular, HAsO₄²⁻ and CrO₄²⁻ are good oxyanion analogues of MoO₄²⁻. Ou et al. [24] proposed a chemisorption mechanism of HAsO₄²⁻ and CrO₄²⁻ onto Mg (OH)₂, either as outer-sphere complexes with a mono-dentate mechanism on defect-poor surface, or multi-dentate inner-sphere mechanisms on defect-rich surface. Their data could be fitted by a Langmuir isotherm. Chemisorption of Sb(III) oxyanion on the surface Mg(OH)₂ would occur according to an outer-sphere complex mechanism that could also be fitted with a Langmuir isotherm [23]. Accordingly, the primary retention mechanism for molybdenum in the present MgO-treated samples would likely be the sorption of molybdate onto brucite, the hydration product of MgO.

Table 8 presents the modeling results for the simplified phase assemblage of key minerals. The modeling confirmed brucite formation in all samples containing MgO, including the 1 % MgO + 3 % LC3 sample, though its low dosage made it undetectable by XRD. The calculated saturation indices of brucite were -0.1 and 0.2 for 1 % and 5 % MgO, respectively and 0.0 for the 1 % MgO + 3 % LC3 (Table S2), which indicating equilibrium with brucite in all cases. Another magnesium-bearing phase, hydrotalcite, was predicted to form based on
<figure><img src="image_12.png" /><figcaption>**Fig. 9.** (a) Acid neutralization capacity (ANC) and (b), (c), and (d) effect of pH on the leaching of Mg, sulfate, and Mo according to the leaching standard US EPA 1313. Red continuous lines represent quantification limits when reached.</figcaption></figure>

the modeling results, but its low concentration likely made it undetected by XRD. Brucite remained the dominant phase after cement addition, with the predicted amount at 1.3 % for brucite and only 0.1 % for hydrotalcite in the 1 % MgO + 3 % LC3 treatment (Table 8). The limited formation of hydrotalcite is further discussed in Section 4.2.

The high point of zero charge (pH<sub>pzc</sub>) of brucite, ranging between 11 and 12 [50], results in a positively charged surface below this threshold, which favors molybdate sorption. In this study, sorption remains effective even after cement addition, despite the increase in pH. However, the results suggest that OPC and Ca(OH)<sub>2</sub> were slightly less effective in retaining molybdate compared to LC3 and SC. This difference can be attributed to the higher equilibrium pH of 12.7 observed with OPC and Ca(OH)<sub>2</sub>, compared to 12.2 with LC3 and SC (Fig. 5), which reduces molybdate sorption by increasing the negative charge on brucite.

Furthermore, the amount of molybdenum leached appears to be primarily governed by the quantity of MgO added to the system, with minimal influence from the amount of cement added (Fig. 2, Fig. 5 and Fig. 7). As illustrated in Fig. 3, the Mo release is strongly dependent on the percentage of MgO added. The relationship between the leached molybdate and the MgO content is linear at low MgO contents, but it becomes non-linear as the MgO content increases. The latter arises from the limited molybdate concentration in solution (0.27 mg/L, Section 3.2.2). As molybdate availability decreases, increasing the amount of the sorbent (brucite) does not result in a proportional increase in molybdate

sorption. The sorption capacity is ultimately constrained by the low molybdate concentration in the solution, as adsorption depends on solution concentration and reaches equilibrium, beyond which additional sorption sites have little effect.

The kinetics of MgO hydration to brucite at room temperature were studied in the literature [50,51]. Their findings revealed that hydration occurs within the first 100 hours (i.e. 3.3 days) of the treatment with the majority of hydration taking place within 24 h. Therefore, this suggests that molybdate could be adsorbed onto the surface shortly after the formation of Mg(OH)<sub>2</sub>. Similarly, Xu et al. [23] showed that Sb was adsorbed by Mg(OH)<sub>2</sub> and reached equilibrium after only 4 h. These fast kinetics could explain the effective Mo retention observed after 7 days of curing for all treatments containing MgO (Fig. 2, Fig. 5 and Fig. 7) and the rapid retention in Mo observed after 1 % MgO + 3 % LC3 addition (Fig. 8). Notably, molybdenum release was already reduced by 67 % after only 6 hours of curing, further supporting the hypothesis that adsorption occurs rapidly after MgO hydration.

Finally, regarding pH sensitivity, a significant reduction of Mo release was observed in the pH range above 9 (Fig. 9-D) compared to the untreated soil. This finding aligns with the hypothesis that Mo retention was due to sorption on brucite. Conversely, the progressive increase in Mo release in the pH range from 10 to 9 and below can be attributed to the gradual dissolution of brucite [23]. Fig. 10 demonstrates that the minimum Mo release at pH > 10 correlates with brucite being close to
<figure><img src="image_13.png" /><figcaption>**Fig. 10.** Saturation indices of brucite (left axis, dark blue) and Mo release (right axis, red) depending on pH value of the 1313 leaching test eluates of soil treated with 1 % MgO + 3 % LC3. SI between -0.5 and 0.5 indicates thermodynamic equilibrium of the mineral.</figcaption></figure>

**Table 8**

Estimation of the simplified mineral phase assemblage for the untreated soil and various treatment additions. All calculations were done at thermodynamic equilibrium, except for the last column, which assumes that only 75 % of the clinker and 50 % of the metakaolin had reacted after 7 days [48,49].

<table><thead><tr><th>Content<br>[wt% of<br>solid mass]</th><th>Untreated<br>soil</th><th>+ 5 %<br>MgO</th><th>+ 5 %<br>LC3</th><th>+ 5 %<br>MgO<br>+ 5 %<br>LC3</th><th>+ 1 %<br>MgO<br>+ 3 %<br>LC3</th><th>+ 1 %<br>MgO<br>+ 3 %<br>LC3<br>(7<br>days)</th></tr></thead><tbody><tr><td>Calcite</td><td>34.0</td><td>31.7</td><td>32.9</td><td>31.0</td><td>33.1</td><td>33.4</td></tr><tr><td>Dolomite</td><td>50.9</td><td>47.5</td><td>48.3<sup>(*)</sup></td><td>45.2<sup>(*)</sup></td><td>48.4<sup>(*)</sup></td><td>48.8<sup>(*)</sup></td></tr><tr><td>Gypsum</td><td>0.9</td><td>0.9</td><td>0.0</td><td>0.0</td><td>0.0</td><td>0.3</td></tr><tr><td>Quartz</td><td>8.0<sup>(*)</sup></td><td>7.5<sup>(*)</sup></td><td>7.6<sup>(*)</sup></td><td>7.1<sup>(*)</sup></td><td>7.6<sup>(*)</sup></td><td>7.7<sup>(*)</sup></td></tr><tr><td>Other minor<br>phases<sup>(**)</sup></td><td>6.2<sup>(*)</sup></td><td>5.8<sup>(*)</sup></td><td>5.9<sup>(*)</sup></td><td>5.6<sup>(*)</sup></td><td>5.9<sup>(*)</sup></td><td>6.0<sup>(*)</sup></td></tr><tr><td>Brucite</td><td>0.0</td><td>6.7</td><td>0.0</td><td>6.1</td><td>1.3</td><td>1.4</td></tr><tr><td>Hydrotalcite</td><td>0.0</td><td>0.0</td><td>0.0</td><td>0.6</td><td>0.1</td><td>0.0</td></tr><tr><td>C-S-H</td><td>0.0</td><td>0.0</td><td>0.3<br>CSH<br>1.1</td><td>1.6<br>CSH<br>1.5</td><td>0.1<br>CSH<br>1.2</td><td>0.7<br>CSH<br>1.5</td></tr><tr><td></td><td></td><td></td><td>1.7<br>CSH<br>1.2</td><td>0.6<br>CSH<br>1.6</td><td>1.2<br>CSH<br>1.3</td><td>0.3<br>CSH<br>1.6</td></tr><tr><td>Portlandite</td><td>0.0</td><td>0.0</td><td>0.0</td><td>0.0</td><td>0.0</td><td>0.0</td></tr><tr><td>Monocarbo<br>-alumin.</td><td>0.0</td><td>0.0</td><td>0.9</td><td>0.0</td><td>0.0</td><td>0.0</td></tr><tr><td>Monosulfo<br>-alum.</td><td>0.0</td><td>0.0</td><td>0.0</td><td>0.0</td><td>0.0</td><td>0.0</td></tr><tr><td>Ettringite</td><td>0.0</td><td>0.0</td><td>2.4</td><td>2.2</td><td>2.3</td><td>1.5</td></tr></tbody></table>

thermodynamic equilibrium with the solution in the pH range 10.5 – 13. The SIs become increasingly negative when the pH decreases from 10.5 to 7, indicating the dissolution of brucite by acid addition and the subsequent release of Mo.

**4.2. Ettringite's role in the immobilization of sulfate and molybdate**

In the preceding section, the sorption of molybdate oxyanion on the brucite surface has been identified as the main retention mechanism.

However, the substitution of sulfate by molybdate in ettringite can be a secondary retention mechanism of molybdenum in the treatments with addition of cements (LC3, OPC, and SC). Indeed, the similar tetrahedral structures of sulfate and molybdate anions facilitates such substitution [28,30].

Ettringite formation was confirmed in all cement-containing samples by XRD (Fig. 6), as well as by phase assemblage modeling (Table 8) and SI calculations (Table S2). The high sulfate content from gypsum (Table 1) consumed most of the aluminum for ettringite formation, limiting the formation of hydrotalcite, monosulfoaluminate, and monocarboaluminate. Modeling predicted gypsum disappearance, no formation of monosulfoaluminate or monocarboaluminate, and small amounts of hydrotalcite (0.6 % for 5 % MgO + 5 % LC3, and 0.1 % for 1 % MgO + 3 % LC3), compared to higher ettringite content (2.2 % and 2.3 %, respectively). A small amount of monocarboaluminate was predicted for LC3 alone, as no competition with hydrotalcite occurred. In the presence of LC3, thermodynamic equilibrium modeling predicted partial dissolution of the Mg-carbonate (dolomite) from the soil and the formation of significant amounts of hydrotalcite, with no ettringite formation. For the 1 % MgO + 3 % LC3 mixture, the calculated pH was 11.7 without dolomite dissolution and 9.9 if dolomite dissolved (compared to 11.8 measured experimentally). The absence of ettringite and the low pH of 9.9 were inconsistent with the experimental data. Dolomite dissolution, which occurs much more slowly than gypsum dissolution, was not observed in the batch tests.

The reduction in sulfate release was observed after addition of 5 wt% cementitious materials to the untreated soil (Fig. 5), indicating its stabilization in ettringite. The greatest decrease was obtained with LC3 by approximately a factor of 3 (from 3200 mg/kg DM to 1100 mg/kg DM). The reduction was also observed with the blends Mg – LC3 blends (Fig. 7). Specifically, the higher the LC3 content is, the higher the ettringite content is, and the lower the sulfate leachate concentration is. Additionally, the slight decrease in molybdenum release with the increased addition of LC3 to the same percentage of MgO (Fig. 7) may also be due to substitution into ettringite.

The kinetics of ettringite precipitation could be influenced by various factors and may proceed slowly, requiring days for completion [30,54]. The progressive formation of ettringite over time could explain the gradual decrease of sulfate release in Fig. 8-C. This could also explain the evolution of molybdenum release in Fig. 8-A. While most of the retention can still be attributed to sorption on brucite, it may be complemented by the progressive formation of ettringite and the substitution of molybdate within it.

Finally, regarding pH sensitivity, the ettringite stability domain is maximal from 11 to 12.5 [30]. Beyond pH 12.5, ettringite was partly dissolved in the present study and led to a high release of sulfate as shown in Fig. 9-C. A similar behavior was observed for molybdenum release (Fig. 9-D). It could be attributed to both the release of the molybdate contained in ettringite and the increase of the negative surface charge of brucite (pH>pH<sub>pzc</sub>).

**4.3. Suitability of the S/S treatment for industrial applications**

**4.3.1. Balance between technical and sustainability issues**

Fig. 2, Fig. 5 and Fig. 7 illustrate the intricate balance between reagent additions necessary to adhere to threshold values. As depicted in Fig. 2 and Fig. 3, the addition of small percentage of MgO proved to be sufficiently effective to reduce molybdenum concentration below the European inert limit threshold value. Therefore, the treatment could be suitable for both disposal in landfills and civil engineering scenarios from a Mo chemical stabilization perspective only.

However, the inclusion of the cementitious material LC3 alongside MgO was deemed essential to enhance the geotechnical properties of the treated soil (Table 7). This is particularly important for disposal scenarios, such as landfilling, or potential valorization applications like road construction [52]. According to French standards, the geotechnical

<footnote>(*) Non-reactive phase. (**) Anorthite, celestite, goethite, palygorskite.</footnote>
durability of the treated soil is considered satisfactory when the ratio of initial CBR (iCBR) to immediate bearing index (IBI) exceeds 1. Even though this ratio was already above 1 before treatment, it was significantly improved by the addition of reagents. The most notable improvements after the addition of 3% of LC3 was a UCS value above 0.35 MPa, the minimal UCS threshold recommended for S/S matrixes by the US EPA [55]. The improvement in mechanical properties was primarily attributed to the formation of C-S-H, a common product of LC3 hydration [49]. Due to its amorphous nature, C-S-H could not be detected by XRD, however, modeling predicted the formation of C-S-H with Ca/Si ratios of 1.1 and 1.3 after the addition of 1 % MgO + 3 % LC3 (0.1 and 1.2 wt% respectively) as reported in Table 8. As expected, a higher pH corresponds to a higher Ca/Si ratio. Notably, no portlandite formation was predicted by assemblage modeling and SI calculation (Table S2). It is consistent with an equilibrium pH of 11–12 (Fig. 8) below the stability domain of portlandite.

LC3 addition also offers some benefits in terms of chemical stabilization since it improves sulfate retention, as evidenced in Fig. 5 and Fig. 7. However, more than 5% LC3 would be required to be below the European inert threshold value on sulfates. The combination of 1 % MgO and 3 % LC3 was also chosen based on cost and sustainability, making it suitable for industrial applications. Indeed, LC3 demonstrates significant potential, enabling the substitution of up to 50 % of clinker and reducing CO₂ emissions by 30 % compared to OPC [56].

### 4.3.2. Fast processing and long-term durability of the treatment

Molybdenum release reached the EU inert threshold value (0.05 mg/L) within just 24 hours (Fig. 8-A). The rapid effectiveness of a treatment is a critical criterion for its viability in field-scale industrial application. Furthermore, the effectiveness of this treatment improves over time and stabilizes significantly below the threshold value.

Another crucial criterion for realistic industrial applications is the durability of the S/S treatment versus pH change. Indeed, Fig. 9-D shows that the release of molybdenum was significantly reduced within the pH range of 8–13, but not for neutral to acidic pH levels. Carbonation by atmospheric CO₂ gas of the treated soil will decrease the pH on the long-term [57], which may destabilize brucite (main sorbent agent for molybdenum) into Mg carbonates and release Mo in the environment. The decrease of pH may also be driven by carboxylic acid as acetic acid by mixing with organic wastes. In complement to solidification purposes, the addition of LC3 cementitious materials will enhance the durability of brucite by creating a supplementary alkaline pH buffer. In the present case, as illustrated in Fig. 9-A, the quantity of acid required to lower the pH below 6 was very substantial (>0.9 kg HNO₃ per kg of DM) due to the high proportion of calcite and dolomite of the untreated soil. This pH-buffering effect of cement and carbonated phases suggests that the treatment's efficiency for molybdenum retention could be maintained under field conditions.

## 5. Conclusion and perspectives

This study evaluated the use of MgO and various types of cements (OPC, SC and LC3) for the S/S of molybdenum-contaminated limestone tunnel sludge, yielding the following key findings:

* Effective stabilization of molybdenum was achieved with small additions of MgO (1–5 wt%), either alone or in combination with cements.

* Brucite ($Mg(OH)_2$) was identified as the primary mechanism for molybdate immobilization via sorption. XRD analysis and geochemical modeling confirmed brucite formation across all formulations.

* Although the addition of cements did not significantly enhance molybdenum immobilization, it was essential for achieving structural integrity reducing sulfate release through ettringite precipitation.

* The combination of 1 wt% MgO and 3 wt% LC3 provides a suitable balance between stabilization and solidification treatment, reduced cost, and a lower carbon footprint achieved through the use of the low-CO₂ cement LC3. However, greater LC3 contents would be necessary to meet the standards for applications requiring high geotechnical performance, such as road construction.

* Performance and chemical durability assessments based on US and EU leaching standards demonstrated high Mo retention under alkaline conditions (pH>11) and after a short curing time (24 hours), with improvements observed over time. However, retention decreased at neutral to acidic pH (pH<8) due to brucite dissolution, although the LC3 cement and soil's natural buffering capacity helped mitigate this effect.

* Future work should focus on assessing long-term durability through dynamic leaching tests, such as EPA methods 1314 and 1315, and evaluating sensitivity to carbonation.

## CRediT authorship contribution statement

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

## 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 acknowledge the reviewers for their relevant comments that greatly helped to improve the manuscript. They also thank the managers and technical/research staff of the analytical team at Holcim Innovation Center and at the Civil and Environmental Engineering Department at Vanderbilt University for their analytical assistance and valuable insights.

## Appendix A. Supporting information

Supplementary data associated with this article can be found in the online version at doi:10.1016/j.jhazmat.2025.138056.

## Data availability

Data will be made available on request.

## References

[1] Smedley, P.L., Kinniburgh, D.G., 2017. Molybdenum in natural waters: a review of occurrence, distributions and controls. Appl Geochem 84, 387–432. https://doi.org/10.1016/j.apgeochem.2017.05.008.

[2] Xu, N., Braida, W., Christodoulatos, C., Chen, J., 2013. A review of molybdenum adsorption in soils/bed sediments: speciation, mechanism, and model applications. Soil Sediment Contam: Int J 22, 912–929. https://doi.org/10.1080/15320383.2013.770438.

[3] Sun, W., Selim, H.M., 2020. Fate and transport of molybdenum in soils: kinetic modeling. Advances in Agronomy. Elsevier, pp. 51–92. https://doi.org/10.1016/bs.agron.2020.06.002.
[4] Cornelis, G., Johnson, C.A., Gerven, T.V., Vandecasteele, C., 2008. Leaching mechanisms of oxyanionic metalloid and metal species in alkaline solid wastes: a review. Appl Geochem 23, 955-976. https://doi.org/10.1016/j.apgeochem.2008.02.001.

[5] Tang, H., Hassan, M.U., Nawaz, M., Yang, W., Liu, Y., Yang, B., 2023. A review on sources of soil antimony pollution and recent progress on remediation of antimony polluted soils. Ecotoxicon Environ Saf 266, 115583. https://doi.org/10.1016/j.ecoenv.2023.115583.

[6] Rahman, S., Rahman, I.M.M., Hasegawa, H., 2023. Management of arsenic-contaminated excavated soils: A review. J Environ Manag 346, 118943. https://doi.org/10.1016/j.jenvman.2023.118943.

[7] Kumar, R., Jing, C., Yan, L., 2024. A critical review on arsenic and antimony adsorption and transformation on mineral facets. S100107424200024X J Environ Sci. https://doi.org/10.1016/j.jes.2024.01.016.

[8] Liu, J., Wu, D., Tan, X., Yu, P., Xu, L., 2023. Review of the Interactions between conventional cementitious materials and heavy metal ions in stabilization/ solidification processing. Materials 16, 3444. https://doi.org/10.3390/ma16093444.

[9] Chen, L., Nakamura, K., Hama, T., 2023. Review on stabilization/solidification methods and mechanism of heavy metals based on OPC-based binders. J Environ Manag 332, 117362. https://doi.org/10.1016/j.jenvman.2023.117362.

[10] Bao, Z., Feng, H., Tu, W., Li, L., Li, Q., 2022. Method and mechanism of chromium removal from soil: a systematic review. Environ Sci Pollut Res. https://doi.org/10.1007/s11356-022-19452-z.

[11] Vollpracht, A., Brameshuber, W., 2016. Binding and leaching of trace elements in Portland cement pastes. Cem Concr Res 79, 76-92. https://doi.org/10.1016/j.cemconres.2015.08.002.

[12] Brandely, M., Coussy, S., Blanc-Biscarat, D., Gourdon, R., Blanck, G., 2022. Chemical stabilization used to reduce geogenic selenium, molybdenum, sulfates and fluorides mobility in rocks and soils from the parisan basin. Environments. https://doi.org/10.3390/environments9070078.

[13] OJEQ. European Decision 2003/33/EC. Council Decision of 19 December 2002 Establishing Criteria and Procedures for the Acceptance of Waste at Landfills Pursuant to Article 16 of and Annex II to Directive 1999/31/EC. 2003, (n.d.). (http://eur-lex.europa.eu/legal-content/EN/TXT/?uri=CELEX%3A2003D0033&qid=1656310258356) (accessed May 27, 2024).

[14] OJEQ. Council Directive 1999/31/EC of 26 April 1999 on the Landfill of Waste. 1999, (n.d.). (https://eur-lex.europa.eu/legal-content/EN/TXT/?uri=CELEX%3A01999L0031-20180704) (accessed May 27, 2024).

[15] AFNOR, NF EN 12457-2 - Characterization of waste - Leaching - Compliance test for leaching of granular waste materials and sludges - Part 2: one stage batch test at a liquid to solid ratio of 10 l/kg for materials with particle size below 4 mm (without or with size reduction), (2002). (https://www.boutique.afnor.org/fr-fr/norme/nf-en-124572/characterisation-des-dechets-livviation-essai-de-conformite-pour-livviation/fa104282/20764) (accessed May 27, 2024).

[16] US EPA, SW-846 Test Method 1311: Toxicity Characteristic Leaching Procedure (TCLP), (1992). (https://www.epa.gov/hw-sw846/sw-846-test-method-1311-toxicity-characteristic-leaching-procedure).

[17] US EPA, PART 268 - Land disposal restrictions, Title 40 -Protection of Environment, Chapter I -Environmental Protection Agency Subchapter I -Solid Wastes, (2024). https://www.ecfr.gov/currenstitle40/chapter-I/subchapter-I/part-268#268.48) (accessed January 5, 2025).

[18] Quintero-Payan, A.C., Huete-Hernández, S., Aguilar-Pozo, V.B., Astals, S., Chimenos, J.M., 2024. Stabilization of metal and metalloids from contaminated soils using magnesia-based tundish deskulling waste from continuous steel casting. Chemosphere 348, 140750. https://doi.org/10.1016/j.chemosphere.2023.140750.

[19] Giro-Palomà, J., Formosa, J., Chimenos, J.M., 2020. Stabilization study of a contaminated soil with metal(loid)s adding different low-grade MgO degrees. Sustainability 12, 7340. https://doi.org/10.3390/su12187340.

[20] Pantazopoulou, E., Ntinoudi, E., Zouboulis, A.I., Mitrakas, M., Yiannoulakis, H., Zampetakis, Th, 2020. Heavy metal stabilization of industrial solid wastes using low-grade magnesia, Portland and magnesia cements. J Mater Cycles Waste Manag 22, 975-985. https://doi.org/10.1007/s10163-020-00985-9.

[21] Zhang, Y., Lu, X., Yu, R., Li, J., Miao, J., Wang, F., 2022. Long-term leachability of Sb in smelting residue stabilized by reactive magnesia under accelerated exposure to strong acid rain. J Environ Manag 301, 113840. https://doi.org/10.1016/j.jenvman.2021.113840.

[22] Suzuki, T., Sue, K., Morotomi, H., Niinae, M., Yokohima, M., Nakata, H., 2019. Immobilization of selenium(VI) in artificially contaminated kaolinite using ferrous ion salt and magnesium oxide. J Environ Chem Eng 7, 102802. https://doi.org/10.1016/j.jece.2018.11.046.

[23] Xu, S., Zhong, Z., Liu, W., Deng, H., Lin, Z., 2020. Removal of Sb(III) from wastewater by magnesium oxide and the related mechanisms. Environ Res 186, 109489. https://doi.org/10.1016/j.envres.2020.109489.

[24] Ou, X., Liu, X., Liu, W., Rong, W., Li, J., Lin, Z., 2018. Surface defects enhance the adsorption affinity and selectivity of Mg(OH)2 towards As(V) and Cr(VI) oxyanions: a combined theoretical and experimental study. Environ Sci: Nano 5, 2570-2578. https://doi.org/10.1039/C8EN00654G.

[25] Wang, F., Wang, H., Jin, F., Al-Tabaa, A., 2015. The performance of blended conventional and novel binders in the in-situ stabilisation/solidification of a contaminated site soil. J Hazard Mater 285, 46-52. https://doi.org/10.1016/j.jhazmat.2014.11.002.

[26] Cubukcuoglu, B., Ouki, S.K., 2019. Comparison of mechanical and leaching behaviours of pulverised fuel ash/low-grade magnesium oxide-cement blended stabilised/solidified baghouse dust. Eur J Environ Civ Eng 23, 771-788. https://doi.org/10.1080/19648189.2017.1310140.

[27] Duan, Y., Liu, X., Zheng, L., Khalid, Z., Long, L., Jiang, X., 2023. MgO-based binders with different formulations for solidifying Pb and Cd in MSWI fly ash: solidification effect and related mechanisms. Process Saf Environ Prot 175, 160-167. https://doi.org/10.1016/j.psep.2023.05.029.

[28] Diaz Castelles, L., Roosz, C., Hot, J., Blotevogel, S., Cyr, M., 2021. Immobilization of molybdenum by alternative cementitious binders and synthetic C-S-H: an experimental and numerical study. Sci Total Environ 789, 148069. https://doi.org/10.1016/j.scitotenv.2021.148069.

[29] Grambow, B., López-García, M., Olmeda, J., Grivé, M., Marty, N.C.M., Regangeon, S., et al., 2020. Retention and diffusion of radioactive and toxic species on cementitious systems: main outcome of the CEBAMA project. Appl Geochem 112, 104480. https://doi.org/10.1016/j.apgeochem.2019.104480.

[30] Chrysochoou, M., Dermatas, D., 2006. Evaluation of ettringite and hydrocalumite formation for heavy metal immobilization: Literature review and experimental study. J Hazard Mater 136, 20-33. https://doi.org/10.1016/j.jhazmat.2005.11.008.

[31] Chen, Q.Y., Tyrer, M., Hills, C.D., Yang, X.M., Carey, P., 2009. Immobilisation of heavy metal in cement-based solidification/stabilisation: A review. Waste Manag 29, 390-403. https://doi.org/10.1016/j.wasman.2008.01.019.

[32] Low Carbon Stabilization and Solidification of Hazardous Wastes, Elsevier, 2022. https://doi.org/10.1016/C2020-0-01157-6.

[33] Sharma, M., Bishnoi, S., Martirena, F., Scrivener, K., 2021. Limestone calcined clay cement and concrete: A state-of-the-art review. Cem Concr Res 149, 106564. https://doi.org/10.1016/j.cemconres.2021.106564.

[34] Scrivener, K., Martirena, F., Bishnoi, S., Maity, S., 2018. Calcined clay limestone cements (LC3). Cem Concr Res 114, 49-56. https://doi.org/10.1016/j.cemconres.2017.08.017.

[35] Reddy, V.A., Solanki, C.H., Kumar, S., Reddy, K.R., Du, Y.-J., 2022. Comparison of limestone calcined clay cement and ordinary Portland cement for stabilization/ solidification of Pb-Zn smelter residue. Environ Sci Pollut Res 29, 11393-11404. https://doi.org/10.1007/s11356-021-16421-w.

[36] Reddy, V.A., Solanki, C.H., Kumar, S., Reddy, K.R., Du, Y.-J., 2020. Pb-Zn smelter residue (LZSR) stabilized using low-carbon, low-cost limestone-calcined clay cement: leachability, chemical speciation, strength, and microstructure. J Hazard Toxic Radioact Waste 24, 04020054. https://doi.org/10.1061/(ASCE)HZ.2153-5515.0000552.

[37] Reddy, V.A., Solanki, C.H., Kumar, S., Reddy, K.R., Du, Y.-J., 2020. Stabilization/ solidification of zinc- and lead-contaminated soil using limestone calcined clay cement (LC3): an environmentally friendly alternative. Sustainability 12, 3725. https://doi.org/10.3390/su12093725.

[38] Wu, H., Song, H., Sun, X., Bi, Y., Fu, S., Yang, N., 2023. Geo-environmental properties and microstructural characteristics of sustainable limestone calcined clay cement (LC3) binder treated Zn-contaminated soils. J Zhejiang Univ Sci A 24, 898-911. https://doi.org/10.1631/jusz.A2200531.

[39] Gu, Y., Li, J.-L., Peng, J.-K., Xing, F., Long, W.-J., Khayat, KamalH., 2020. Immobilization of hazardous ferronickel slag treated using ternary limestone calcined clay cement. Constr Build Mater 250, 118837. https://doi.org/10.1016/j.conbuildmat.2020.118837.

[40] Afnor, N.F. P94-093 - Soils: Investigation and testing - Determination of the compaction characteristics of a soil - Standard Proctor test - Modified Proctor test, (2014). (https://www.boutique.afnor.org/fr-fr/norme/nf-p94093/sols-reconnaissance-et-essais-determination-des-references-de-compactage-du/fa185491/43924) (accessed May 27, 2024).

[41] Brandely, M., Coussy, S., Blanc-Biscarat, D., Gourdon, R., Blanck, G., 2022. Chemical stabilization used to reduce geogenic selenium, molybdenum, sulfates and fluorides mobility in rocks and soils from the Parisian basin. Environments. https://doi.org/10.3390/environments9070078.

[42] US EPA, SW-846 method 1313 liquid-solid partitioning as a function of extract ph using a parallel batch extraction procedure, (2012). (https://www.epa.gov/sites/default/files/2015-12/documents/1313.pdf).

[43] AFNOR, NF EN 13286-47 - Unbound and hydraulically bound mixtures - Part 47: Test method for the determination of California bearing ratio, immediate bearing index and linear swelling, (2021). (https://www.boutique.afnor.org/fr-fr/norme/nf-en-1328647/melanges-traites-et-melanges-non-traites-aux-liants-hydrauliques-partie-47/-fa199701/335006).

[44] AFNOR, NF EN 13286-41 - Unbound and hydraulically bound mixtures - Part 41: Test method for the determination of the compressive strength of hydraulically bound mixtures, (2021). (https://www.boutique.afnor.org/fr-fr/norme/nf-en-1328641/melanges-traites-et-melanges-non-traites-aux-liants-hydrauliques-partie-41/-fa199700/334998).

[45] van der Lee, J., De Windt, L., Lagneau, V., Goblet, P., 2003. Module-oriented modeling of reactive transport with HYTEC. Geosciences 265-275. https://doi.org/10.1016/S0098-3004(03)00004-9.

[46] Blanc, Ph, Lassin, A., Piantone, P., Azaroual, M., Jacquemet, N., Fabbri, A., et al., 2012. Thermodmod: A geochemical database focused on low temperature water/ rock interactions and waste materials. Appl Geochem 27, 2107-2116. https://doi.org/10.1016/j.apgeochem.2012.06.002.

[47] Appelo, C.A.J., Postma, D., 2004. Balkema Publishers (NL). Geochem. Groundw Pollut.

[48] Avet, F., Scrivener, K., 2018. Investigation of the calcined kaolinite content on the hydration of Limestone Calcined Clay Cement (LC3). Cem Concr Res 107, 124-135. https://doi.org/10.1016/j.cemconres.2018.02.016.

[49] Zunino, F., Dhandapani, Y., Ben Haha, M., Skibsted, J., Joseph, S., Krishnan, S., et al., 2022. Hydration and mixture design of calcined clay blended cements: review by the RILEM TC 282-CCL. Mater Struct 55, 234. https://doi.org/10.1617/s11527-022-02060-1.
[50] Pettauer, M., Baldermann, A., Eder, S., Dietzel, M., 2024. Hydration of MgO: reaction kinetics and pH control on brucite crystal morphology. *acs.cgd.4c00243* Cryst Growth Des. https://doi.org/10.1021/acs.cgd.4c00243.

[51] Thomas, J.J., Musso, S., Prestini, I., 2014. Kinetics and Activation Energy of Magnesium Oxide Hydration. J Am Ceram Soc 97, 275–282. https://doi.org/10.1111/jace.12661.

[52] Shah, V., Scott, A., 2021. Use of kaolinite clays in development of a low carbon MgO-clay binder system. Cem Concr Res 144, 106422. https://doi.org/10.1016/j.cemconres.2021.106422.

[53] Tian, Q., Guo, B., Chuaicham, C., Sasaki, K., 2020. Mechanism analysis of selenium (VI) immobilization using alkaline-earth metal oxides and ferrous salt. Chemosphere 248, 126123. https://doi.org/10.1016/j.chemosphere.2020.126123.

[54] Xu, L., Wang, P., Zhang, G., 2012. Formation of ettringite in Portland cement/calcium aluminate cement/calcium sulfate ternary system hydrates at lower temperatures. Constr Build Mater 31, 347–352. https://doi.org/10.1016/j.conbuildmat.2011.12.078.

[55] Goodarzi, A.R., Movahedrad, M., 2017. Stabilization/solidification of zinc-contaminated kaolin clay using ground granulated blast-furnace slag and different types of activators. Appl Geochem 81, 155–165. https://doi.org/10.1016/j.apgeochem.2017.04.014.

[56] Vizcaíno-Andrés, L.M., Sánchez-Berriel, S., Damas-Carrera, S., Pérez-Hernández, A., Scrivener, K.L., Martirena-Hernández, J.F., 2015. Industrial trial to produce a low clinker, low carbon cement. Mater Construcc 65, e045. https://doi.org/10.3989/mc.2015.00614.

[57] Li, M., Cai, G., Wang, Q., Liu, S., He, H., Liu, X., et al., 2023. The state of the art of carbonation technology in geotechnical engineering: a comprehensive review. Renew Sustain Energy Rev 171, 112986. https://doi.org/10.1016/j.rser.2022.112986.