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Article

Inhibition Mechanism of Calcium Hydroxide on Arsenic Volatilization During Sintering of Contaminated Excavated Soils

by
Xu Li
1,2,
Yu Jin
1,
Yaocheng Wang
2,3,4,
Zhijun Dong
1 and
Weipeng Feng
1,*
1
Institute of Technology for Future Industry, School of Science and Technology Instrument Application Engineering, Shenzhen University of Information Technology, Shenzhen 518172, China
2
College of Civil and Transportation Engineering, Shenzhen University, Shenzhen 518060, China
3
Key Laboratory for Resilient Infrastructures of Coastal Cities (MOE), College of Civil and Transportation Engineering, Shenzhen University, Shenzhen 518060, China
4
Guangdong Provincial Key Laboratory of Durability for Marine Civil Engineering, Shenzhen University, Shenzhen 518060, China
*
Author to whom correspondence should be addressed.
Sustainability 2025, 17(20), 9027; https://doi.org/10.3390/su17209027 (registering DOI)
Submission received: 11 September 2025 / Revised: 7 October 2025 / Accepted: 9 October 2025 / Published: 12 October 2025
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Abstract

Urbanization generates large quantities of arsenic-contaminated excavated soils that pose environmental risks due to arsenic volatilization during high-temperature sintering processes. While these soils have potential for recycling into construction materials, their reuse is hindered by arsenic release. This study demonstrated calcium hydroxide (Ca(OH)2) as a highly effective additive for suppressing arsenic volatilization during soil sintering, while simultaneously improving material properties. Through comprehensive characterization using inductively coupled plasma-mass spectrometry (ICP-MS), scanning electron microscopy (SEM) and X-ray microtomography (μCT), energy dispersive spectroscopy (EDS), X-ray diffraction (XRD), Fourier Transform Infrared Spectroscopy (FTIR) and X-ray photoelectron spectroscopy (XPS), results demonstrated that Ca(OH)2 addition (0.5–2 wt.%) reduces arsenic volatilization by 57% through formation of thermally stable calcium arsenate (Ca3(AsO4)2). Ca(OH)2 acted via two mechanisms: (a) chemical immobilization through Ca-As-O compound formation, (b) physical encapsulation in a calcium-aluminosilicate matrix during liquid-phase sintering, and (c) pH buffering that maintains arsenic in less volatile forms. Optimal performance was achieved at 0.5% Ca(OH)2, yielding 9.14 MPa compressive strength (29% increase) with minimal arsenic leaching (<110 ppb). Microstructural analysis showed Ca(OH)2 promoted densification while higher doses increased porosity. This work provides a practical solution for safe reuse of arsenic-contaminated soils, addressing both environmental concerns and material performance requirements for construction applications.

1. Introduction

Excavated soils, classified as construction and demolition waste (CDW), are primarily generated from urban infrastructure development, subsurface construction, and land expansion projects [1]. In recent years, with rapid urbanization, underground spaces have become critical for transport networks, utility systems, and commercial facilities [2], resulting in excavated soil production surpassing 1.5 billion tons annually in China and 180 million tons in Japan [3,4] which accounted for 30–40% and circa 20% of their total solid waste [5,6], respectively. Notably, the Guangdong-Hong Kong-Macao Greater Bay Area (GBA) alone contributes circa 283 million tons annually to China’s total [7], a particularly pressing issue as Guangdong province simultaneously faces some of the nation’s most severe arsenic contamination in soils [8]. Arsenic contamination in soils originates from both natural processes (geogenic weathering, volcanic/hydrothermal activities, atmospheric deposition) and anthropogenic activities (mining/smelting, industrial emissions, agricultural/urban practices), with contributions varying by geological setting and land-use history [9]. Such massive quantities exert severe pressure on landfill capacity while posing significant contamination risks [10,11,12]. Therefore, to align with circular economy goals, developing efficient recycling strategies for excavated soils is no longer optional but a necessity for sustainable construction.
Different from naturally deposited clays, these anthropogenic soils typically exhibit heterogeneity in composition, particle sizes and possible contamination. Specifically, through on-site size-based segregation, the coarse fraction demonstrates inherent suitability for construction aggregates (e.g., recycled aggregates, manufactured sand) [13], whereas fine particles are alternatively valorized as sustainable alternatives in soil-cement composites and brick manufacturing. Several studies have demonstrated that properly processed excavated soils present chemical composition comparable to clay components and can be sintered into bricks, ceramic tiles, and lightweight aggregates [14,15,16,17], alleviating the traditional exclusive reliance on the clay resource. The sintering process leverages high-temperature treatment to induce densification and phase transformations in soil particles. The thermal treatment essentially triggers physicochemical changes in clay minerals (the naturally formed component predominant in clay) [18], proceeding through distinct temperature-dependent stages. When the temperature is low (≤400 °C), the major reaction stems from the evaporation of water which was physically absorbed on the external surfaces. At intermediate temperatures (400–650 °C), clay minerals such as kaolinite with the ideal formula of Al4[(OH)8Si4O10] undergo dehydroxylation, losing structural water to form metakaolinite (Al2O3·SiO2) [19]. Upon further heating (950–1000 °C), metastable phases recrystallize into γ-Al2O3 spinel phase or Al-Si spinel phase from aluminosilicate precursors, with the formation of amorphous SiO2 accompanied [20], noting that the above sintering mechanisms usually involve reaction of several phases other than only one clay mineral at high temperature. Cations like K+, Na+, Fe2+, Fe3+, Ca2+, and Mg2+ in natural clay minerals can serve as fluxing agents during sintering, facilitating liquid-phase formation by reducing viscosity and lowering the eutectic temperature [21]. Thus, the soil materials are densified through particle rearrangement and viscous flow, ultimately intensifying firing shrinkage and developing mechanical strength as bulk materials driven by accelerated pore collapse and mass transport.
However, soils contaminated either by anthropogenic or geogenic arsenic due to deep excavation pose significant challenges to their sustainable utilization. The two primary oxidation states that exist in soil are AsIII (arsenite) and AsV (arsenate), which are mostly toxic, with AsIII exhibiting higher toxicity (50-fold), mobility and reactivity but lower stability and abundance compared to AsV in the soil environment [4,22,23]. Their toxicity, volatility and bio-accumulation in the environment will cause severe health effects, including skin lesions, cardiovascular diseases, and various cancers [24]. Such concerns are particularly exacerbated during high-temperature processes such as sintering, where elevated temperatures promote arsenic volatilization from contaminated soils. Current research indicates that the volatilization rate of arsenic increases proportionally with temperature [25,26]. Thermogravimetric analysis of coals over the range of 600–1500 °C revealed two distinct mass loss peaks for arsenic release [27], with three critical temperature zones identified: <600 °C, 600–1000 °C, and >1000 °C [22]. The temperature-dependent volatilization of arsenic directly overlaps with the sintering window of excavated soils for construction materials, creating a fundamental challenge for recycling arsenic-contaminated soils.
To inhibit arsenic volatilization during high-temperature processes, researchers in coal combustion fields have explored the interaction between arsenic and various metallic elements, with calcium demonstrating particularly effective arsenic retention capabilities [28,29]. X-ray diffraction results indicated that Ca3(AsO4)2 and Ca3As2O7 were formed when volatilized arsenic reacted with CaO [30]. Hydrated lime (Ca(OH)2) was found to excel for Al2O3, SiO2 and kaolinite in arsenic sorption at 600 °C in adsorption experiments using simulated flue gas containing gaseous As2O3(g), ultimately forming stable Ca3(AsO4)2 precipitates [31], which were reported to be stable until exposed to temperatures above 1450 °C [32]. Up to now, regarding the As-contaminated soil, previous studies have demonstrated that lime, metal oxides (e.g., Fe, Mn, Al, and Ca), ordinary Portland cement, etc., are effective for the solidification/stabilization (S/S) treatment [33,34,35]. When cement and calcium hydroxide were used for S/S treatment, arsenic leachability was effectively reduced, attributing to the formation of calcite, precipitates, and other insoluble calcium-arsenic compounds (e.g., Ca3(AsO4)2, Ca3(AsO3)2, CaAsO3(OH)·2H2O, and CaHAsO4, CaH4(AsO4)2, Ca(H2AsO4)2) [36]. The toxicity of arsenic was also reduced by the oxidation reaction, where AsIII was converted to AsV, promoted by cementitious materials [37]. Research revealed that sintered bricks produced from arsenic-containing sludge [38], copper slag flotation tailings [39], etc., exhibited satisfactory quality, with arsenic leaching concentrations remaining below permissible environmental regulatory limits. However, the arsenic volatilization during the sintering process was rarely considered in these studies, raising environmental safety concerns regarding this utilization method. Therefore, while calcium-based additives have demonstrated effectiveness in absorbing arsenic volatilization during coal combustion as well as in S/S treatment in As-contaminated soils, their application in sintering As-contaminated soils for construction materials requires deeper investigation.
To this end, this study aims to investigate the effect of calcium hydroxide on arsenic solidification and the microstructure of sintered excavated soil. The arsenic solidification effect was evaluated before and after sintering via inductively coupled plasma-mass spectrometry (ICP-MS). X-ray microtomography (μCT) and scanning electron microscopy (SEM) were used to characterize the microstructure of sintered samples. By combining the results from X-ray diffraction (XRD), energy dispersive spectroscopy (EDS), X-ray photoelectron spectroscopy (XPS), and Fourier transform infrared spectroscopy (FTIR) of Ca(OH)2-treated excavated soil before and after sintering, a fundamental mechanism for Ca(OH)2 on sintered excavated soil was discussed in depth.

2. Materials and Methods

2.1. Raw Materials

The investigated soil was collected from an excavated site in Shenzhen. As shown in Figure 1a, the raw soil was first oven-dried at 105 °C until constant weight was achieved. The dried soil was then ground using a mechanical grinder (Model BJ-800A, BAIJIE CO. Ltd., Shenzhen, China) and sieved through a 150 μm mesh. The resulting processed soil (Figure 1b) was used for subsequent experiments. Water used in this study was ultrapure water produced by a Milli-Q system. Ca(OH)2 (analytical grade) was purchased from Macklin Scientific Corp (Shanghai, China).
The concentration of arsenic in the soil was 38.3 ± 2.6 ppm, measured by Inductively Coupled Plasma-Mass Spectrometry (ICP-MS) according to the Chinese National Standard HJ 803-2016 (Soil and sediment-determination of aqua regia extracts of 12 metal elements-Inductively coupled plasma mass spectrometry) [40]. The chemical composition of the processed soil is shown in Table 1, where SiO2 and Al2O3 collectively account for over 90 wt.% of the soil composition. Particle size distribution is presented in Figure 2. Figure 3a displays the mineral composition of the processed soil. The result indicates that the predominant mineral phases in the excavated soil are quartz (SiO2), kaolinite (Al4[(OH)8Si4O10]), and muscovite (KAl2(AlSi3O10)(OH)2). Thermal behavior of the soil shown in Figure 3b revealed several distinct events, including the moisture evaporation below 200 °C, dehydroxylation of kaolinite to form metakaolin at 500–700 °C (peak at 600 °C) and structural stabilization above 900 °C.

2.2. Sample Preparation

The sample mixing design is shown in Table 2. The liquid-to-solid ratios (L/S) were fixed to 0.3 for all the samples. Sample preparation followed the protocol schematized in Figure 4. The solid mixture (soil and Ca(OH)2) was initially homogenized using a handheld homogenizer (DDQ-B02L1) for 120 s to ensure uniform dispersion. Subsequently, ultrapure water was introduced while maintaining mixing for an additional 150 s under consistent conditions. The homogenized slurry was then hermetically sealed in a zip-lock bag and aged for 3 days (72 h) under controlled laboratory conditions (20 ± 2 °C). Following the aging period, the stabilized mixtures were compacted into a cylindrical geometry (h × Ø: 15 mm × 10 mm) using a hardened steel die and digitally controlled electro-hydraulic testing system (EH-8305Y, Wance Group, Shenzhen, China). The dry-pressing protocol involves axial loading at a constant displacement rate of 1.0 ± 0.1 mm/min until reaching 2.0 ± 0.2 MPa, followed by a pressure maintenance for 30 ± 1 s under closed-loop control. The compacted specimens were conditioned at ambient laboratory temperature (20 ± 2 °C, RH 50 ± 5%) for 48 h to facilitate moisture equilibration. Subsequent thermal treatment followed the programmed sintering profile illustrated in Figure 5, concluding with furnace-cooling to room temperature (20 ± 2 °C).
After subjecting them to the compressive test, the fractured residues were collected and ground using an agate mortar and pestle, followed by particle size fractionation through a 75 μm sieve. The resulting homogenized powder was allocated for the following test including ICP-MS, XRD, FTIR, SEM-EDS and XPS.

2.3. Characterization Methods

2.3.1. Compression Strength Test

Compressive strength was determined per GB/T 17671-2021 (Test method for cement mortar strength-ISO method) [41] using a digitally controlled electro-hydraulic testing system (EH-8305Y, Wance Group). Cylindrical specimens (h × Ø: 15 mm × 10 mm) were axially loaded at a constant displacement rate of 2.00 mm/min until failure. Sixteen samples were tested for each Ca(OH)2 dosage group to ensure statistical reliability of the compressive strength measurements.

2.3.2. ICP-MS

The arsenic concentration was quantitatively determined for processed soil, Ca(OH)2 and all specimens before and after sintering following Chinese National Standard HJ 803-2016 (Soil and sediment-determination of aqua regia extracts of 12 metal elements-Inductively coupled plasma mass spectrometry). For analysis, 100 mg of homogenized soil samples were digested following standard aqua regia extraction procedures before measurement using inductively coupled plasma mass spectrometry (ICP-MS, iCAP RQ, Thermo Scientific, Waltham, MA, USA).
Toxicity characteristic leaching tests were conducted under systematically controlled pH conditions (acidic/neutral/alkaline) by mixing 2.00 g of homogenized powder with three leaching media: 0.01 M HCl (pH = 2.00), ultrapure water (Milli-Q, 18.2 MΩ·cm, pH = 5.60), and 0.01 M NaOH (pH = 12.00). The liquid-to-solid ratios were fixed at 10:1. The mixtures were shaken by an oscillator initially to ensure particle suspension. Following oscillation, samples underwent quiescent equilibration for 24 h prior to 0.45 μm PVDF membrane filtration. The leached arsenic quantification was determined by ICP-MS.
Three replicate samples were analyzed for each condition to ensure data reproducibility.

2.3.3. X-Ray Diffraction

Crystalline phase identification was performed using a Bruker D8 Advance diffractometer (Bruker Corp., Billerica, MA, USA) equipped with a Cu-Kα radiation source (λ = 1.5406 Å) operating at 40 kV and 40 mA (1.6 kW power). Measurements were conducted in Bragg–Brentano θ–2θ geometry with a 0.5° divergence slit, employing a continuous scan mode across the 5–70° 2θ range at a scanning rate of 1°/min with 0.02° step resolution. Three replicate measurements were conducted for each sample to confirm phase identification consistency.

2.3.4. Fourier Transform Infrared Spectroscopy

Molecular bonding analysis was performed using a Bruker INVENIOS FTIR spectrometer. Samples were homogenized with AR-grade KBr (1:100 sample-to-KBr mass ratio) and pressed into 7 mm diameter pellets under 10-ton hydraulic pressure. Spectra were acquired in transmission mode (64 scans at 4 cm−1 resolution) across the mid-infrared range (4000–400 cm−1). Triplicate measurements were obtained for each sample to verify spectral reproducibility.

2.3.5. X-Ray Photoelectron Spectroscopy

Chemical states of arsenic, carbon, and calcium were investigated using a Thermo Scientific ESCALAB 250Xi+ spectrometer with a monochromatic Al Kα X-ray source. Powder samples were uniformly dispersed on indium foil and degassed under ultrahigh vacuum at 120 °C for 12 h to minimize surface contamination. High-resolution spectra of As3d and C1s core levels were acquired with a hemispherical analyzer (pass energy = 20 eV, step size = 0.05 eV) under dual-beam charge compensation. Binding energies were calibrated against adventitious carbon (C1s at 284.8 eV), and spectral fitting was performed using Voigt peak profiles (70% Gaussian, 30% Lorentzian) with Shirley background subtraction. Triplicate measurements were obtained for each sample to verify spectral reproducibility.

2.3.6. Scanning Electron Microscope

The microstructural and elemental distribution of the samples was investigated using a ZEISS Crossbeam 350 (Carl Zeiss AG, Oberkochen, Germany) focused ion beam scanning electron microscope (FIB-SEM) system equipped with an Oxford Instruments Ultim Max silicon drift detector (SDD) for energy-dispersive X-ray spectroscopy (EDS). High-resolution secondary electron imaging was performed at 3 kV accelerating voltage to preserve fine structural details, while EDS elemental mapping (Si Kα, Al Kα, Ca Kα, and As Lα lines) was conducted at 15 kV with a dwell time of 50 μs/pixel.

2.3.7. X-Ray Microtomography (μCT)

The three-dimensional microstructure of all samples was characterized using an Eclipse XRM 3D Microscope system (PrismaXRM-800, SIGRAY INC., Concord, CA, USA). One representative sample from each group was subjected to μCT scanning to obtain comprehensive 3D microstructural data. The scans were performed at 150 keV tube energy and 8.0 W power, with a 6.2× magnification achieved by setting an object-detector distance of 258.98 mm and the source-object distance of 41.86 mm, resulting in a pixel resolution of 8.0 µm. Cylindrical specimens (h × Ø: 15 mm × 10 mm) were scanned through 360° rotation with a 0.2° interval, capturing a total of 1801 projection images at 0.3 s exposure per frame (2940 × 2340 pixel resolution).
Raw projection data underwent reconstruction via the Filter Backprojection algorithm [42] to generate volumetric datasets. After reconstruction, data analysis was conducted via Avizo (version 2019.4, Thermo Fisher Scientific Inc.) as shown in Figure 6. To optimize computational efficiency and minimize edge artifacts [43], analysis was restricted to a volume of interest (VOI) of 8003 voxels, corresponding to a physical dimension of 6.43 mm3. Each voxel stored 16-bit grayscale values (ranging from 0 to 65,536), where the grayscale intensity correlated with the local material density and atomic number [44,45]. This allowed for clear discrimination between the pore phase (black, low X-ray absorption) and the solid phase (bright, high X-ray absorption). Furthermore, variations in grayscale intensity within the solid phase revealed compositional heterogeneity, with denser or higher-atomic-number components appearing brighter due to stronger X-ray attenuation. Before segmentation, median filtering was applied to reduce noise [46]. The grayscale images were then binarized using Otsu’s method [47] to separate pores from the solid matrix. As this study focused on pore network analysis, the solid-phase components were not further subdivided. The pore morphology was quantified in terms of equivalent diameter and volume. The equivalent diameter, calculated using Equation (1), represents the diameter of a sphere with equivalent volume to the pore, providing a standardized measure of pore size.
E q D i a m e t e r = 6 × V o l u m e 3 d π 3
where V o l u m e 3 d denotes the volume of a pore.

3. Results

3.1. Compressive Strength of Sintered Excavated Soils

The compressive strength results of sintered excavated soils with varying Ca(OH)2 additions reveal a discernible trend based on additive content as shown in Figure 7, though variations exist due to sample heterogeneity. The control group (CA0%) exhibits moderate strength with an average of 7.40 MPa, while the CA0.5% samples demonstrate the highest performance in compression at 9.14 MPa, suggesting that a small Ca(OH)2 addition enhances densification and bonding during sintering [48]. In contrast, increasing Ca(OH)2 to 1% leads to a notable decline in strength (6.98 MPa), likely due to excessive Ca(OH)2 interfering with particle consolidation or forming weaker phases like pores. The CA2% group shows a partial recovery to 7.40 MPa, though still underperforming compared to CA0.5%, indicating that higher Ca(OH)2 content may partially compensate for but cannot fully restore optimal mechanical properties.
A closer examination of individual measurements reveals variability within each group, particularly in CA0% and CA0.5%, where strengths range from 4 to 10 MPa and 7–10 MPa, respectively. This suggests that while CA0.5% generally improves consistency, some samples in CA0% achieve comparable peak strength, possibly due to natural soil heterogeneity. The CA1% and CA2% groups display tighter ranges (5–8 MPa), implying that higher Ca(OH)2 additions stabilize strength but at a lower overall level. Notably, CA0.5% not only achieves the highest mean but also the most consistently high values, with 75% of samples exceeding 8 MPa, whereas CA1% and CA2% rarely surpass this threshold.
When comparing trends, the non-linear response to Ca(OH)2 dosage highlights an optimal threshold near 0.5%. The initial strength increase with CA0.5% may stem from improved particle bonding, whereas higher Ca(OH)2 levels likely introduce porosity or unreacted phases that weaken the matrix.

3.2. Arsenic Behavior Analysis

The arsenic concentration data (Figure 8a) reveal distinct trends before and after sintering. The processed soil initially contained 38,322 ppm As, while pure Ca(OH)2 exhibited a significantly lower arsenic content (298 ppm), confirming that Ca(OH)2 itself was not a source of arsenic contamination. The addition of Ca(OH)2 to the excavated soil prior to sintering reduced the measured arsenic concentration compared to the original soil, suggesting that Ca(OH)2 contributed to arsenic immobilization [45]. After sintering, the sample without Ca(OH)2 (CA0%) showed a substantial decrease in arsenic content (from 27,125 ppm to 12,281 ppm), indicating significant arsenic volatilization at high temperatures. In contrast, samples with Ca(OH)2 addition (CA0.5%, CA1%, CA2%) retained higher arsenic concentrations after sintering, demonstrating that Ca(OH)2 effectively suppressed arsenic volatilization during thermal treatment.
The leaching behavior of arsenic (Figure 8b) varied depending on both Ca(OH)2 content and the pH of the leaching medium. In the absence of Ca(OH)2 (CA0%), the leached arsenic concentrations were similar across acidic (78 ppb), neutral (70 ppb), and alkaline (104 ppb) conditions, with only a slight increase in the alkaline medium. However, when Ca(OH)2 was added, distinct leaching patterns emerged. Under acidic conditions (pH = 2), the CA0.5% sample exhibited a sharp increase in arsenic leaching (1239 ppb), which gradually decreased with higher Ca(OH)2 content (933 ppb for CA1%, 874 ppb for CA2%). A similar but less pronounced trend was observed in neutral conditions (pH = 7), where leaching peaked at 646 ppb for CA0.5% and declined to 295 ppb for CA2%. In contrast, alkaline conditions (pH = 12) maintained consistently low arsenic leaching (95–109 ppb) regardless of Ca(OH)2 content.
These findings suggest that while Ca(OH)2 enhances arsenic retention during sintering, the resulting immobilized arsenic phases remain susceptible to dissolution in acidic environments. The decreasing leaching trend with increasing Ca(OH)2 content implies that excess Ca(OH)2 raises the medium pH, thereby reducing arsenic mobility. This is supported by the minimal leaching observed in alkaline conditions, where high pH further stabilizes arsenic within the sintered matrix. The data collectively indicate that Ca(OH)2 modifies both the thermal stability and leaching behavior of arsenic, with optimal immobilization achieved at higher Ca(OH)2 dosages under neutral to alkaline conditions.

3.3. Microstructural and Elemental Analysis

3.3.1. SEM-EDS

The microstructural evolution of excavated soils with varying Ca(OH)2 content before and after sintering was investigated using SEM-EDS, as illustrated in Figure 9. In the unsintered state, the CA0% (Figure 9(a1–a3)) sample exhibited a dense and uniform structure with minimal porosity, characterized by well-defined kaolinite lamellar stacking [49] and blocky quartz grains. The distinct boundaries and smooth surfaces suggested minimal interaction between the mineral phases (quartz and kaolinite), with no observable cementation. In contrast, the CA2% sample before sintering displayed a rougher texture due to the presence of Ca(OH)2, which absorbed water and induced localized swelling [50], forming initial pores and micro-agglomerates. Deposits, likely calcium arsenate (Ca3(AsO4)2) precipitates resulting from Ca(OH)2-As(V) interactions, were also observed, though the underlying kaolinite and quartz morphology remained discernible.
After sintering, the CA0% (Figure 9(b1–b3)) sample retained much of its original morphology, with only slight particle rounding at the edges due to partial melting. A limited amount of glassy phase, formed from the melting of aluminosilicates at high temperatures, was present, though the microstructure remained predominantly crystalline. In comparison, the CA2% (Figure 9(d1–d3)) sample exhibited significant coarsening, with particles covered by a more extensive vitrified layer. Both sintered samples (CA0% and CA2%) contained needle-like phases (Figure 9(b1,d3)), attributed to the formation of mullite (3Al2O3·2SiO2) under high-temperature conditions [51]. The abundance of SiO2 (59.19%) and Al2O3 (30.77%) in the raw soil (Table 1) facilitated this reaction. The mullite needles in CA2% were notably coarser, likely due to the decomposition of Ca(OH)2 into CaO during sintering. Acting as a fluxing agent, CaO reduced the crystallization temperature [52], thereby promoting lateral crystal growth of mullite.
The lamellar kaolinite structures observed in the unsintered samples (Figure 9a,c) were not distinct after sintering, though their outlines remained visible with reduced edge sharpness. The original loose particle arrangements were replaced by denser agglomerates, indicating consolidation and partial vitrification during the sintering process.
The elemental distribution of Si, Al, Ca, and As in excavated soils with varying Ca(OH)2 content before and after sintering was analyzed through SEM-EDS mapping (Figure 10 and Figure 11). All samples exhibited a dominant Si-Al matrix with trace amounts of As. In CA0% and CA0.5%, the Ca signal appeared sporadically due to its low concentration. Additionally, the strong signals from the Si-Al matrix likely overshadowed the weak Ca peaks [53]. In contrast, CA1% and CA2% displayed clear Ca enrichment owing to the deliberate addition of Ca(OH)2. Before sintering (Figure 10c,d), Ca was primarily adsorbed or present in an amorphous state along the edges of kaolinite layers and within the pores of Si-Al gels. After sintering (Figure 11), the enhanced detectability of Ca in EDS analysis can be seen, as the newly formed crystalline or glassy phases provided a more uniform distribution of Ca.
The compositional trends were further illustrated in the CaO-SiO2-Al2O3 ternary diagram (Figure 12). Given the high initial proportions of SiO2 and Al2O3 in the excavated soil, the data points clustered near the SiO2-Al2O3 edge, with minimal CaO contribution (0.74% in CA0%). The addition of Ca(OH)2 shifted the composition slightly toward the CaO edge, though the overall change before and after sintering was not pronounced. This observation suggests that while Ca(OH)2 incorporation influenced the local chemistry, the bulk composition remained dominated by the original Si-Al matrix.

3.3.2. μCT

As shown in Figure 13, the pore phase segmented from the reconstructed data was visualized via binarization as the value was set to 1. The real pore 3D morphology and the arbitrary pore model by equivalent radius are illustrated in Figure 13a and Figure 13b, respectively. The segmented pore phase obtained from μCT analysis reveals distinct trends in pore morphology with increasing Ca(OH)2 content. While the control sample (CA0%) shows fine, uniformly distributed pores, CA0.5% exhibits moderate porosity with isolated micropores that enhance strength. Higher Ca(OH)2 additions (1–2%) lead to significantly increased pore sizes and interconnectivity, forming macroporous networks that compromise mechanical performance despite maintaining arsenic immobilization.
As shown in Figure 14a, the equivalent diameter distribution shifts toward larger values with Ca(OH)2 addition, particularly for CA1% and CA2% samples. For instance, pores with diameters >50 µm become significantly more abundant in CA1% (1203 counts) and CA2% (2091 counts) compared to CA0% (928 counts) and CA0.5% (1429 counts). This trend is corroborated by the pore volume distribution (Figure 14c), where CA1% and CA2% exhibit a marked increase in large pores (e.g., volumes > 1 × 107 µm3), while CA0% and CA0.5% are dominated by smaller pores (<5 × 106 µm3). Notably, the CA0.5% sample demonstrates an intermediate pore structure, with a moderate increase in equivalent diameter (e.g., 31.4 µm vs. 27.5 µm in CA0% for the third size bin) but no significant shift toward extremely large pores. In contrast, CA1% and CA2% show a bimodal distribution, with a secondary peak emerging for pores > 100 µm, coinciding with a 3–5× rise in large-pore volume frequency. This divergence highlights a threshold effect: below 0.5% Ca(OH)2, pore growth is limited, whereas beyond 1% Ca(OH)2, excessive pore coarsening occurs, likely due to altered sintering dynamics or gas entrapment.
The heat map analysis of pore equivalent diameter (Figure 14b) and volume (Figure 14d) further elucidates the pore structure evolution. In CA0%, the pore size distribution is concentrated within a narrow range of 50–450 µm, indicating a fine and relatively uniform pore network. In contrast, CA1% and CA2% exhibit a significantly broader distribution, extending up to 750–850 µm, suggesting substantial pore coarsening. This shift toward larger pore sizes implies that Ca(OH)2 addition promotes pore coalescence, leading to the formation of enlarged voids or even microcracks during sintering. A similar trend is observed in the pore volume distribution (Figure 14d), where CA0% remains confined to smaller volumes, while CA1% and CA2% display a pronounced expansion into higher volume ranges, indicating that Ca(OH)2 incorporation disrupts the fine-scale porosity of the calcined soil and the development of a coarser, more heterogeneous pore structure.
These findings align with the compressive strength trends, where CA0.5% achieved optimal strength (9.14 MPa) due to balanced densification, while CA1% and CA2% suffered strength losses (6–7 MPa) from excessive porosity. The μCT data directly explain this mechanical decline: the proliferation of large pores (>50 µm) in CA1% and CA2% (Figure 14a) creates stress-concentration sites [54,55], while their higher total pore volumes (evident in Figure 14b) reduce load-bearing capacity. Thus, the non-linear relationship between Ca(OH)2 content, pore structure, and strength underscores the critical role of pore morphology in governing mechanical performance.

3.4. Phase Composition and Chemical Speciation Analysis

3.4.1. X-Ray Diffraction

As shown in Figure 15, phase transformations can be identified during thermal treatment. For the raw soil (Figure 15a), the dominant crystalline phases were identified as quartz (SiO2), kaolinite (Al4[(OH)8Si4O10]), and muscovite (KAl2(AlSi3O10)(OH)2), consistent with typical clay mineral compositions. Upon calcination, significant structural changes were observed. The complete disappearance of kaolinite’s characteristic diffraction peaks (particularly at ~12.4° and 24.9° 2θ) confirms its thermal decomposition, which occurs through a two-stage process: a) dehydroxylation between 400 and 650 °C [18], converting the crystalline structure into metastable amorphous metakaolin (Al2Si2O7), and b) subsequent transformation into spinel-type phases before ultimately forming mullite (3Al2O3·2SiO2) at higher temperatures [56]. Notably, while thermodynamic calculations predict mullite formation above 1000 °C, its characteristic peaks (e.g., ~16.4°, 26.0°, and 40.8° 2θ) were absent in our XRD patterns (Figure 15b). This absence can be attributed to the relatively low crystallization degree of mullite formed at 1000 °C, as previous studies [21] have demonstrated that well-defined mullite diffraction only becomes detectable after extended heating above 1100 °C.
The XRD analysis did not detect any crystalline arsenic-containing phases due to the combination of arsenic’s low concentration below 5 wt.% (under the XRD detection limit [35]) and its probable incorporation into amorphous phases or surface adsorption. The observed zinc oxide peaks at approximately 31.8°, 34.4°, and 36.3° 2θ served as internal reference markers for instrument calibration without affecting phase identification.

3.4.2. Fourier Transform Infrared Spectroscopy

The infrared spectroscopy analysis was conducted to characterize the atomic bonds in the excavated soil samples, with measurements performed in the range of 4000–400 cm−1 at a resolution of 2 cm−1 with 40 scans. As shown in Figure 16a, the uncalcined soil exhibited distinct absorption bands in several characteristic regions. In the high-frequency region (3700–3600 cm−1), two prominent absorption peaks at 3694 cm−1 and 3650 cm−1 were observed, corresponding to the stretching vibrations of hydroxyl groups. The bands at 3700 and 3624 cm−1 can be specifically attributed to O-H vibrations in kaolinite [57,58], which can be distinguished from illite/smectite based on their characteristic vibration patterns in the 3750–3400 cm−1 range.
The mid-frequency region revealed several important vibrational modes associated with carbonate formation. The absorption bands in the 2200–2400 cm−1 range, corresponding to the antisymmetric stretching vibration of CO2, appeared following Ca(OH)2 addition during the aging process. This observation indicates that Ca(OH)2 absorbed atmospheric CO2 under humid conditions, forming calcium carbonate (CaCO3) or intermediate calcium bicarbonate (Ca(HCO3)2) products [59]. Simultaneously, the characteristic peak at 1420 cm−1, attributed to the asymmetric C-O stretching vibration of carbonate groups (CO32−), further confirmed the presence of CaCO3 [58,60,61]. Notably, the intensity of these carbonate-related vibration peaks progressively increased with higher Ca(OH)2 addition (CA0%–CA2%), demonstrating that the carbonation reaction was enhanced by greater Ca(OH)2 dosage. The aging process (as mentioned in Figure 4) played a crucial role in this transformation, where the alkalinity of Ca(OH)2 facilitated CO2 absorption and carbonation reactions, while trace moisture present in the soil (such as interlayer water in kaolinite) provided the necessary reaction medium [62]. These findings indicate that increasing Ca(OH)2 content led to more pronounced carbonation and greater formation of carbonate species in the system. The spectral region between 1250 and 1000 cm−1 displayed multiple silicate-related vibrations, including Si-O stretching at 1114 cm−1 and 1160 cm−1 [63], Si-O-Si stretching at 1033 cm−1, and another Si-O stretching mode at 1004 cm−1. Additional characteristic peaks were observed at lower frequencies, including an Al-OH bending vibration at 912 cm−1 [59] and 830 cm−1 [63], Si-O and Si-O-Si stretching vibrations at 796, 754, and 690 cm−1, and Si-O-Al bending vibrations at 541, 466, and 430 cm−1 [64]. These results provide comprehensive evidence of the mineralogical composition and structural characteristics of the uncalcined soil samples.
After calcination at 1000 °C (Figure 16b), the infrared spectra of the soil samples exhibited significant structural changes. The disappearance of O-H stretching vibrations at 3700 and 3624 cm−1, along with the Al-OH bending vibration at 912 cm−1, confirms the dehydroxylation of kaolinite and the removal of structural water. This thermal decomposition leads to the destruction of Al-OH bonds and subsequent rearrangement of the Al-O framework [63]. The structural degradation of kaolinite is further evidenced by the broadening of Si-O stretching vibrations at 1114, 1033, and 1004 cm−1, as well as the weakening or disappearance of lattice vibrations at 790, 754, and 694 cm−1. Additionally, the emergence of a strong Si-O stretching band at 1073 cm−1 suggests the transformation of kaolinite into metakaolin, accompanied by the loss of OH groups and the amorphization of its crystalline structure [65]. The reduced intensities and broadening of the Si-O-Al bending vibrations at 541, 466, and 430 cm−1 further indicate the weakening of Si-O-Al bonding [62]. Similarly, the diminished peak at 830 cm−1, corresponding to Al-OH bonds, confirms the thermal destruction of the kaolinite crystal structure [55]. These observations collectively demonstrate that high-temperature calcination induces dehydroxylation, structural collapse, and phase transformation in the excavated soil, resulting in an amorphous aluminosilicate network [66].
The most distinct difference among calcined excavated soils with varying Ca(OH)2 contents was observed in the peaks around 2900–3000 cm−1, which corresponds to the C-H stretching vibration of bicarbonate (HCO3−) [67]. This peak formation is attributed to the highly reactive CaO derived from the decomposition of Ca(OH)2 during calcination. Upon cooling, the CaO preferentially reacts with CO2/H2O rather than with aluminosilicate phases, leading to the formation of calcium bicarbonate (Ca(HCO3)2). In contrast, the raw soil (without Ca(OH)2 addition) showed no such peak due to its inherently low calcium content (only 0.74% CaO) and the absence of reactive CaO, which is essential for bicarbonate formation. Additionally, the weakening of the 1420 cm−1 peak (characteristic of CO32−) was observed, as CaCO3 began decomposing above 580 °C and completely transforms into CaO and CO2 at 1000 °C, resulting in the disappearance of carbonate-related vibrational modes. These findings further confirm that the carbonation process is highly dependent on the availability of reactive CaO, which is generated in situ from Ca(OH)2 decomposition during calcination.
Notably, while FTIR is sensitive to As-O bond vibrations depending on functional group specificity, no distinct As-related peaks were observed in the spectra. Pentavalent arsenic (As5+) shows a characteristic absorption band at 800–850 cm−1 corresponding to As-O stretching vibrations [68], as demonstrated by Ca3(AsO4)2 which exhibits a peak at 820 cm−1 [69]. However, in this study, the potential As-O signals were likely obscured by strong overlapping absorption from silicates (Si-O, 1000–1100 cm−1) and aluminates (Al-O) present in the soil matrix. The high content of SiO2 (59.19 wt.%) and Al2O3 (30.77 wt.%), as shown in Table 1, particularly contributed to this signal interference. Additional spectral interference came from moisture in the samples, where O-H stretching vibrations around 3400 cm−1 further complicated the spectral interpretation. These combined matrix effects and inherent detection limitations provide a clear explanation for the absence of identifiable As-O peaks in the FTIR data.

3.4.3. X-Ray Photoelectron Spectroscopy

The chemical state of arsenic in sintered excavated soils with varying Ca(OH)2 content was investigated using XPS analysis, as shown in Figure 17. Due to the extremely low arsenic content in the samples, the obtained spectra exhibited relatively poor signal-to-noise ratios. Nevertheless, the As3d spectra revealed discernible trends in arsenic speciation with increasing Ca(OH)2 addition.
In the CA0% sample, no distinct arsenic peaks could be reliably fitted, indicating arsenic concentrations below the detection limit of the technique. As reported in [70,71,72], the binding energy of 44.5–45.2 indicates the formation of As(V)-O. The CA0.5% sample showed only a very weak As(V) peak at 44.5 eV, suggesting minimal arsenic retention in this composition. However, a clear trend emerged with higher Ca(OH)2 content, where the peak area corresponding to As(V) at 44.5 eV progressively increased in the CA1% and CA2% samples. This observation demonstrates that the addition of Ca(OH)2 enhanced the immobilization of arsenic in its pentavalent state during the sintering process.

4. Discussion

The comprehensive experimental results demonstrate that Ca(OH)2 addition governs arsenic immobilization through synergistic chemical and physical mechanisms during soil sintering. The chemical stabilization pathway initiates during aging, where Ca(OH)2 reacts with arsenate to form calcium arsenate (Ca3(AsO4)2), evidenced by XPS detection of As(V) species and supported by literature demonstrating thermal stability up to 1450 °C [32]. This primary chemical mechanism persists through sintering, where Ca(OH)2 decomposition into CaO [73] creates additional arsenic-binding sites while simultaneously altering the soil matrix through fluxing effects [52]. The fluxing effect is particularly evident in the observed coarsening of mullite needles in CA2% samples (Figure 9(d3)), where increased CaO content enhances crystal growth through improved mass transport. The detectable Ca signals of CA2% in SEM-EDS results (Figure 11(d4)) further indicate localized enrichment and the formation of calcium-aluminosilicate phases. Concurrently, though limited by low As concentrations, the XPS results suggest that Ca(OH)2 enhances the immobilization of As in its pentavalent state, as indicated by the increasing As(V) peak area with higher Ca(OH)2 content. This aligns with the FTIR observations, where the absence of distinct As-O peaks is likely due to the strong overlapping signals from the Si-Al matrix, as is evident by ternary diagram analyses (Figure 12).
The alkaline environment created by Ca(OH)2 addition influences As speciation and mobility. The pH-dependent leaching behavior implies such an additional stabilization mechanism. In alkaline conditions (pH = 12), all samples show minimal As release due to the stability of calcium arsenate phases and strong adsorption on iron/aluminum oxides. The reduced leaching in acidic and neutral media with increasing Ca(OH)2 content demonstrates the system’s buffering capacity, where excess CaO maintains a local pH higher to prevent the complete dissolution of As-bearing phases. The pH-dependent leaching behavior suggests particular suitability for applications in neutral-to-alkaline environments characteristic of urban construction settings. The short-term leaching tests demonstrate effective arsenic immobilization under controlled pH conditions, but long-term stability under environmental exposure (e.g., freeze-thaw cycles, carbonation, or acid rain) remains to be validated. Given that calcium arsenate phases are known to be stable in alkaline to neutral environments [31], the Ca(OH)2 modified sintered matrices may offer inherent resistance to degradation, but accelerated aging tests are needed to confirm this hypothesis. Moreover, while acidic conditions showed slightly elevated arsenic release, the Ca(OH)2-treated materials still maintained leaching concentrations well below China’s regulatory limits (GB 5085.3-2007) [74].
However, the μCT pore quantification reveals how Ca(OH)2-induced microstructural modifications contribute to physical encapsulation. While all Ca(OH)2-treated samples show increased porosity from gas evolution (H2O and CO2 release), the 0.5% Ca(OH)2 specimens develop optimized pore networks characterized by isolated micropores (Figure 14) that correlate with both maximum compressive strength (9.14 MPa) and effective arsenic retention (81%). This microstructure results from balanced liquid phase formation, evidenced by SEM-observed vitrification (Figure 9b) and EDS-detected Ca-Al-Si phase development (Figure 11), which creates diffusion barriers around arsenic compounds [75]. In contrast, higher Ca(OH)2 doses (1–2%) generate excessive interconnected macroporosity visible in μCT reconstructions (Figure 13 and Figure 14), explaining their reduced strength despite maintained arsenic immobilization.
The multi-technique evidence converges to demonstrate Ca(OH)2’s dual stabilization mechanism: chemical fixation through calcium arsenate formation and physical encapsulation within a modified aluminosilicate matrix. The 0.5% Ca(OH)2 dosage emerges as optimal by balancing these mechanisms (sufficient CaO for arsenic binding and controlled viscosity for pore structure management), while avoiding the detrimental porosity effects seen at higher doses. Although the absolute arsenic concentrations may vary geographically, the demonstrated inhibition mechanisms should remain effective across the range of concentrations (10–100 ppm) commonly encountered in arsenic-affected regions worldwide [8]. The fundamental chemical interactions between calcium and arsenic species, along with the physical encapsulation processes, are governed by thermodynamics rather than absolute concentration. This mechanistic framework explains both the environmental and engineering performance of Ca(OH)2-treated sintered soils.

5. Conclusions

This study systematically investigates the role of Ca(OH)2 in controlling arsenic behavior during the sintering of excavated soils with respect to mechanical strength, leaching behavior, pore distribution, and chemical composition by employing compressive tests, leaching experiments via ICP-MS, SEM-EDS, µCT, XRD, FTIR, and XPS. The conclusions of the study are as follows:
(1)
Ca(OH)2 addition significantly reduced arsenic volatilization during sintering, with CA0.5% samples retaining 81% of the initial arsenic content (38.3 ppm) compared to only 51.8% in untreated (CA0%) samples. This stabilization occurs through the formation of thermally stable calcium arsenate (Ca3(AsO4)2) as confirmed by XPS analysis showing As(V) retention in CA1% and CA2% samples.
(2)
Ca(OH)2 acted both chemically, by forming stable As-bearing compounds, and physically, by promoting liquid-phase sintering that encapsulates arsenic. The decomposition of Ca(OH)2 into CaO enhances densification and reduces As mobility, while the alkaline environment further suppresses leaching, maintaining As concentrations below 110 ppb in all Ca(OH)2-modified samples under alkaline conditions.
(3)
An optimal Ca(OH)2 dosage of 0.5% achieves the best balance between arsenic retention and mechanical performance. This formulation yields a compressive strength of 9 MPa (29% higher than CA0%) while minimizing pore connectivity. Higher Ca(OH)2 contents lead to increased porosity and reduced strength, despite continued As immobilization benefits, indicating that excessive Ca(OH)2 addition can compromise structural properties while providing diminishing returns for As stabilization.
The findings present a viable approach for sustainable recycling of As-contaminated excavated soils into construction materials, addressing both environmental and engineering requirements. The method offers a practical solution for hazardous waste management in urban development projects, though further research is needed to evaluate long-term stability under real-world conditions. This work contributes to the development of sustainable construction materials while providing an effective strategy for managing contaminated soils in environmentally sensitive regions.

Author Contributions

Conceptualization, X.L. and Y.J.; Methodology, X.L. and Y.J.; Formal analysis, X.L. and W.F.; Investigation, X.L. and W.F.; Resources, Y.W. and Z.D.; Writing—original draft, X.L. and W.F.; Writing—review & editing, X.L., Y.J. and W.F.; Supervision, Y.J. and W.F.; Funding acquisition, Y.J. and W.F. All authors have read and agreed to the published version of the manuscript.

Funding

The authors gratefully acknowledge the financial supports of the National Nature Science Foundation of China (52408281, 52378251), Shenzhen Science and Technology Program (JCYJ20241202130800001), and Special Team Program for the research of new generation alkali-activated low-carbon cement from SZIIT (TD2024E001).

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

The data presented in this study are available on request from the corresponding author.

Conflicts of Interest

The authors declare no conflicts of interest.

References

  1. Rahman, S.; Rahman, I.M.M.; Hasegawa, H. Management of Arsenic-Contaminated Excavated Soils: A Review. J. Environ. Manag. 2023, 346, 118943. [Google Scholar] [CrossRef]
  2. Kishii, T. Utilization of Underground Space in Japan. Tunn. Undergr. Space Technol. 2016, 55, 320–323. [Google Scholar] [CrossRef]
  3. Huang, T.; Kou, S.; Liu, D.; Li, D.; Xing, F. Evaluation of the Techno-Economic Feasibility for Excavated Soil Recycling in Shenzhen, China. Sustainability 2022, 14, 3028. [Google Scholar] [CrossRef]
  4. Tabelin, C.B.; Igarashi, T.; Villacorte-Tabelin, M.; Park, I.; Opiso, E.M.; Ito, M.; Hiroyoshi, N. Arsenic, Selenium, Boron, Lead, Cadmium, Copper, and Zinc in Naturally Contaminated Rocks: A Review of Their Sources, Modes of Enrichment, Mechanisms of Release, and Mitigation Strategies. Sci. Total Environ. 2018, 645, 1522–1553. [Google Scholar] [CrossRef] [PubMed]
  5. Huang, B.; Wang, X.; Kua, H.; Geng, Y.; Bleischwitz, R.; Ren, J. Construction and Demolition Waste Management in China through the 3R Principle. Resour. Conserv. Recycl. 2018, 129, 36–44. [Google Scholar] [CrossRef]
  6. Lu, W.; Webster, C.; Chen, K.; Zhang, X.; Chen, X. Computational Building Information Modelling for Construction Waste Management: Moving from Rhetoric to Reality. Renew. Sustain. Energy Rev. 2017, 68, 587–595. [Google Scholar] [CrossRef]
  7. Yu, B.; Wang, J.; Wu, H.; Wong, A.B.; Liao, Y.; Zuo, J. Self-Fulfillment Degree of Construction and Demolition Waste Management Capability Based on the Triple-Balance Theory: A Case Study of Guangdong-Hong Kong-Macao Greater Bay Area. Waste Manag. 2021, 133, 99–109. [Google Scholar] [CrossRef] [PubMed]
  8. Zhou, Y.; Niu, L.; Liu, K.; Yin, S.; Liu, W. Arsenic in Agricultural Soils across China: Distribution Pattern, Accumulation Trend, Influencing Factors, and Risk Assessment. Sci. Total Environ. 2018, 616–617, 156–163. [Google Scholar] [CrossRef]
  9. Hettick, B.E.; Cañas-Carrell, J.E.; French, A.D.; Klein, D.M. Arsenic: A Review of the Element’s Toxicity, Plant Interactions, and Potential Methods of Remediation. J. Agric. Food Chem. 2015, 63, 7097–7107. [Google Scholar] [CrossRef]
  10. Mandal, B.K.; Suzuki, K.T. Arsenic Round the World: A Review. Talanta 2002, 58, 201–235. [Google Scholar] [CrossRef]
  11. Li, J.; Kosugi, T.; Riya, S.; Hashimoto, Y.; Hou, H.; Terada, A.; Hosomi, M. Use of Batch Leaching Tests to Quantify Arsenic Release from Excavated Urban Soils with Relatively Low Levels of Arsenic. J. Soils Sediments 2017, 17, 2136–2143. [Google Scholar] [CrossRef]
  12. Smedley, P.L.; Kinniburgh, D.G. A Review of the Source, Behaviour and Distribution of Arsenic in Natural Waters. Appl. Geochem. 2002, 17, 517–568. [Google Scholar] [CrossRef]
  13. Li, Y.; Yi, P.; Du, H.; Liu, W.; Mi, T.; Huang, L.; Gao, X.; Sun, X.; Xing, F. Activation of Locally Excavated Spoil for Utilization in Limestone Calcined Clay Cement (LC3). Constr. Build. Mater. 2024, 420, 135518. [Google Scholar] [CrossRef]
  14. Haas, M.; Galler, R.; Scibile, L.; Benedikt, M. Waste or Valuable Resource—A Critical European Review on Re-Using and Managing Tunnel Excavation Material. Resour. Conserv. Recycl. 2020, 162, 105048. [Google Scholar] [CrossRef]
  15. Li, H.; Zhang, J.; Zheng, W.; Cui, T.; Yang, Y. Study on the Properties and Process Parameters of Different Clays in Disc Granulation. Materials 2020, 13, 1714. [Google Scholar] [CrossRef]
  16. Li, H.; Yang, Y.; Zheng, W.; Chen, L.; Bai, Y. Immobilization of High Concentration Hexavalent Chromium via Core-Shell Structured Lightweight Aggregate: A Promising Soil Remediation Strategy. Chem. Eng. J. 2020, 401, 126044. [Google Scholar] [CrossRef]
  17. Gao, W.; Jian, S.; Li, B.; Li, X.; Tan, H.; Huang, W.; Yang, X. Solidification of Zinc in Lightweight Aggregate Produced from Contaminated Soil. J. Clean. Prod. 2020, 265, 121784. [Google Scholar] [CrossRef]
  18. Martin, R.T.; Bailey, S.W.; Eberl, D.D.; Fanning, D.S.; Guggenheim, S.; Kodama, H.; Pevear, D.R.; Środoń, J.; Wicks, F.J. Report of the Clay Minerals Society Nomenclature Committee: Revised Classification of Clay Materials. Clays Clay Miner. 1991, 39, 333–335. [Google Scholar] [CrossRef]
  19. Cheng, H.; Liu, Q.; Yang, J.; Ma, S.; Frost, R.L. The Thermal Behavior of Kaolinite Intercalation Complexes-A Review. Thermochim. Acta 2012, 545, 1–13. [Google Scholar] [CrossRef]
  20. Ghorbel, A.; Fourati, M.; Bouaziz, J. Microstructural Evolution and Phase Transformation of Different Sintered Kaolins Powder Compacts. Mater. Chem. Phys. 2008, 112, 876–885. [Google Scholar] [CrossRef]
  21. Wang, S.; Gainey, L.; Mackinnon, I.D.R.; Allen, C.; Gu, Y.; Xi, Y. Thermal Behaviors of Clay Minerals as Key Components and Additives for Fired Brick Properties: A Review. J. Build. Eng. 2023, 66, 105802. [Google Scholar] [CrossRef]
  22. Wang, C.; Liu, H.; Zhang, Y.; Zou, C.; Anthony, E.J. Review of Arsenic Behavior during Coal Combustion: Volatilization, Transformation, Emission and Removal Technologies. Prog. Energy Combust. Sci. 2018, 68, 1–28. [Google Scholar] [CrossRef]
  23. Lim, K.T.; Shukor, M.Y.; Wasoh, H. Physical, Chemical, and Biological Methods for the Removal of Arsenic Compounds. BioMed Res. Int. 2014, 2014, 503784. [Google Scholar] [CrossRef]
  24. Tchounwou, P.B.; Yedjou, C.G.; Patlolla, A.K.; Sutton, D.J. Heavy Metal Toxicity and the Environment. In Molecular, Clinical and Environmental Toxicology; Luch, A., Ed.; Experientia Supplementum; Springer: Basel, Switzerland, 2012; Volume 101, pp. 133–164. ISBN 978-3-7643-8339-8. [Google Scholar]
  25. Clarke, L.B. The Fate of Trace Elements during Coal Combustion and Gasification: An Overview. Fuel 1993, 72, 731–736. [Google Scholar] [CrossRef]
  26. Lu, H.; Chen, H.; Li, W.; Li, B. Transformation of Arsenic in Yima Coal during Fluidized-Bed Pyrolysis. Fuel 2004, 83, 645–650. [Google Scholar] [CrossRef]
  27. Liu, H.; Pan, W.-P.; Wang, C.; Zhang, Y. Volatilization of Arsenic During Coal Combustion Based on Isothermal Thermogravimetric Analysis at 600–1500 °C. Energy Fuels 2016, 30, 6790–6798. [Google Scholar] [CrossRef]
  28. Tang, Q.; Liu, G.; Yan, Z.; Sun, R. Distribution and Fate of Environmentally Sensitive Elements (Arsenic, Mercury, Stibium and Selenium) in Coal-Fired Power Plants at Huainan, Anhui, China. Fuel 2012, 95, 334–339. [Google Scholar] [CrossRef]
  29. Sterling, R.; Helble, J. Reaction of Arsenic Vapor Species with Fly Ash Compounds: Kinetics and Speciation of the Reaction with Calcium Silicates. Chemosphere 2003, 51, 1111–1119. [Google Scholar] [CrossRef]
  30. Jadhav, R.A.; Fan, L.S. Capture of Gas-Phase Arsenic Oxide by Lime: Kinetic and Mechanistic Studies. Environ. Sci. Technol. 2001, 35, 794–799. [Google Scholar] [CrossRef]
  31. Mahuli, S.; Agnihotri, R.; Chauk, S.; Ghosh-Dastidar, A.; Fan, L.-S. Mechanism of Arsenic Sorption by Hydrated Lime. Environ. Sci. Technol. 1997, 31, 3226–3231. [Google Scholar] [CrossRef]
  32. Wang, J.; Tomita, A. A Chemistry on the Volatility of Some Trace Elements during Coal Combustion and Pyrolysis. Energy Fuels 2003, 17, 954–960. [Google Scholar] [CrossRef]
  33. Li, Y.; Tang, J.X.; Zhang, T.; Lian, M.H.; Tian, X.Z. Study on Chemical Stabilization in Arsenic Contaminated Soil: A Review. In Progress in Environmental Protection and Processing of Resource; Trans Tech Publications Ltd.: Baech, Switzerland, April 2013; Volume 295, pp. 1089–1092. [Google Scholar]
  34. Tsang, D.C.W.; Yip, A.C.K.; Olds, W.E.; Weber, P.A. Arsenic and Copper Stabilisation in a Contaminated Soil by Coal Fly Ash and Green Waste Compost. Environ. Sci. Pollut. Res. Int. 2014, 21, 10194–10204. [Google Scholar] [CrossRef] [PubMed]
  35. Yoon, I.-H.; Moon, D.H.; Kim, K.-W.; Lee, K.-Y.; Lee, J.-H.; Kim, M.G. Mechanism for the Stabilization/Solidification of Arsenic-Contaminated Soils with Portland Cement and Cement Kiln Dust. J. Environ. Manag. 2010, 91, 2322–2328. [Google Scholar] [CrossRef]
  36. Kundu, S.; Gupta, A.K. Immobilization and Leaching Characteristics of Arsenic from Cement and/or Lime Solidified/Stabilized Spent Adsorbent Containing Arsenic. J. Hazard. Mater. 2008, 153, 434–443. [Google Scholar] [CrossRef]
  37. Coussy, S.; Paktunc, D.; Rose, J.; Benzaazoua, M. Arsenic Speciation in Cemented Paste Backfills and Synthetic Calcium–Silicate–Hydrates. Miner. Eng. 2012, 39, 51–61. [Google Scholar] [CrossRef]
  38. Hassan, K.M.; Fukushi, K.; Turikuzzaman, K.; Moniruzzaman, S.M. Effects of Using Arsenic–Iron Sludge Wastes in Brick Making. Waste Manag. 2014, 34, 1072–1078. [Google Scholar] [CrossRef]
  39. Liu, D.-G.; Min, X.-B.; Ke, Y.; Chai, L.-Y.; Liang, Y.; Li, Y.-C.; Yao, L.-W.; Wang, Z.-B. Co-Treatment of Flotation Waste, Neutralization Sludge, and Arsenic-Containing Gypsum Sludge from Copper Smelting: Solidification/Stabilization of Arsenic and Heavy Metals with Minimal Cement Clinker. Environ. Sci. Pollut. Res. 2018, 25, 7600–7607. [Google Scholar] [CrossRef]
  40. HJ 803-2016; Soil and Sediment-Determination of Aqua Regia Extracts of 12 Metal Elements-Inductively Coupled Plasma Mass Spectrometry. Chinese National Standard: Beijing, China, 2016.
  41. GB/T 17671-2021; Test Method of Cement Mortar Strength (ISO Method). Chinese National Standard: Beijing, China, 2021.
  42. Carmignato, S.; Dewulf, W.; Leach, R. (Eds.) Industrial X-Ray Computed Tomography; Springer International Publishing: Cham, Switzerland, 2018; ISBN 978-3-319-59571-9. [Google Scholar]
  43. Hubler, M.H.; Gelb, J.; Ulm, F.-J. Microtexture Analysis of Gas Shale by XRM Imaging. J. Nanomech. Micromech. 2017, 7, 04017005. [Google Scholar] [CrossRef]
  44. Bullard, J.W.; Hagedorn, J.; Ley, M.T.; Hu, Q.; Griffin, W.; Terrill, J.E. A Critical Comparison of 3D Experiments and Simulations of Tricalcium Silicate Hydration. J. Am. Ceram. Soc. 2018, 101, 1453–1470. [Google Scholar] [CrossRef] [PubMed]
  45. Provis, J.L.; Myers, R.J.; White, C.E.; Rose, V.; van Deventer, J.S.J. X-Ray Microtomography Shows Pore Structure and Tortuosity in Alkali-Activated Binders. Cem. Concr. Res. 2012, 42, 855–864. [Google Scholar] [CrossRef]
  46. Kaestner, A.; Lehmann, E.; Stampanoni, M. Imaging and Image Processing in Porous Media Research. Adv. Water Resour. 2008, 31, 1174–1187. [Google Scholar] [CrossRef]
  47. Otsu, N. A Threshold Selection Method from Gray-Level Histograms. IEEE Trans. Syst. Man Cybern. 1979, 9, 62–66. [Google Scholar] [CrossRef]
  48. Hamano, K.; Muroyama, K.; Nakagawa, Z.; Saito, K. Effects of Addition of Calcium Compounds on Sintering of Magnesia. J. Ceram. Assoc. Jpn. 1979, 87, 474–482. [Google Scholar] [CrossRef]
  49. Fashina, B.; Deng, Y. Stacking Disorder and Reactivity of Kaolinites. Clays Clay Min. 2021, 69, 354–365. [Google Scholar] [CrossRef]
  50. Sang, Y.; Zhao, J.R. Reduction of Water Absorption Capacity of Cellulose Fibres for Its Application in Cementitious Materials. J. Compos. Mater. 2015, 49, 2757–2763. [Google Scholar] [CrossRef]
  51. Sanya, O.T.; Owoeye, S.S.; Isinkaye, O.E.; Simon, O.T. Chemical, Phase and Structural Change of Mullite Synthesized during Sintering of Kaolin. Int. J. Appl. Ceram. Tech. 2020, 17, 2259–2264. [Google Scholar] [CrossRef]
  52. Ding, D.; Jiang, W.; Xiao, G.; Fang, Y.; Zhu, X.; Chong, X.; Jin, E.; Luo, J.; Lei, C. Effects of CaO Addition on the Properties and Microstructure of Low-Grade Bauxite-Based Lightweight Ceramic Proppant. J. Aust. Ceram. Soc. 2023, 59, 957–967. [Google Scholar] [CrossRef]
  53. Newbury, D.E.; Leapman, R.D. Detection Limits for Analytical Electron Microscopy with Electron Energy Loss Spectrometry and Energy-Dispersive x-Ray Spectrometry. Proc. Annu. Meet. Electron. Microsc. Soc. Am. 1993, 51, 594–595. [Google Scholar] [CrossRef]
  54. Feng, W.; Jin, Y.; Zheng, D.; Fang, Y.; Dong, Z.; Cui, H. Study of Triethanolamine on Regulating Early Strength of Fly Ash-Based Chemically Foamed Geopolymer. Cem. Concr. Res. 2022, 162, 107005. [Google Scholar] [CrossRef]
  55. Zhang, Z.; Wang, H. The Pore Characteristics of Geopolymer Foam Concrete and Their Impact on the Compressive Strength and Modulus. Front. Mater. 2016, 3, 38. [Google Scholar] [CrossRef]
  56. Deiters, M.; Rozanski, C.; Schack, T.; Haist, M. Influence of Calcined Clay on Micromechanical Properties and Creep of Hardened Cement Paste. ce papers 2023, 6, 431–440. [Google Scholar] [CrossRef]
  57. Chen, Y.; Furmann, A.; Mastalerz, M.; Schimmelmann, A. Quantitative Analysis of Shales by KBr-FTIR and Micro-FTIR. Fuel 2014, 116, 538–549. [Google Scholar] [CrossRef]
  58. Ho, G.D.; Tabelin, C.B.; Tangviroon, P.; Tamamura, S.; Igarashi, T. Effects of Cement Addition on Arsenic Leaching from Soils Excavated from Projects Employing Shield-Tunneling Method. Geoderma 2021, 385, 114896. [Google Scholar] [CrossRef]
  59. Amar, M.; Ladduri, B.; Alloul, A.; Benzerzour, M.; Abriak, N.-E. Microstructure and Mechanical Properties of Geopolymers Utilizing Excavated Soils, Metakaolin and Slags. J. Build. Eng. 2024, 86, 108755. [Google Scholar] [CrossRef]
  60. Tabelin, C.B.; Sasaki, R.; Igarashi, T.; Park, I.; Tamoto, S.; Arima, T.; Ito, M.; Hiroyoshi, N. Simultaneous Leaching of Arsenite, Arsenate, Selenite and Selenate, and Their Migration in Tunnel-Excavated Sedimentary Rocks: I. Column Experiments under Intermittent and Unsaturated Flow. Chemosphere 2017, 186, 558–569. [Google Scholar] [CrossRef] [PubMed]
  61. Andersen, F.A.; Brečević, L.; Beuter, G.; Dell’Amico, D.B.; Calderazzo, F.; Bjerrum, N.J.; Underhill, A.E. Infrared Spectra of Amorphous and Crystalline Calcium Carbonate. Acta Chem. Scand. 1991, 45, 1018–1024. [Google Scholar] [CrossRef]
  62. Castellano, M.; Turturro, A.; Riani, P.; Montanari, T.; Finocchio, E.; Ramis, G.; Busca, G. Bulk and Surface Properties of Commercial Kaolins. Appl. Clay Sci. 2010, 48, 446–454. [Google Scholar] [CrossRef]
  63. Vizcayno, C.; de Gutiérrez, R.M.; Castello, R.; Rodriguez, E.; Guerrero, C.E. Pozzolan Obtained by Mechanochemical and Thermal Treatments of Kaolin. Appl. Clay Sci. 2010, 49, 405–413. [Google Scholar] [CrossRef]
  64. González-García, D.M.; Téllez-Jurado, L.; Jiménez-Álvarez, F.J.; Zarazua-Villalobos, L.; Balmori-Ramírez, H. Evolution of a Natural Pozzolan-Based Geopolymer Alkalized in the Presence of Sodium or Potassium Silicate/Hydroxide Solution. Constr. Build. Mater. 2022, 321, 126305. [Google Scholar] [CrossRef]
  65. Cheng, S.; Ge, K.; Sun, T.; Shui, Z.; Chen, X.; Lu, J.-X. Pozzolanic Activity of Mechanochemically and Thermally Activated Coal-Series Kaolin in Cement-Based Materials. Constr. Build. Mater. 2021, 299, 123972. [Google Scholar] [CrossRef]
  66. Frost, R.L.; Makó, É.; Kristóf, J.; Kloprogge, J.T. Modification of Kaolinite Surfaces through Mechanochemical Treatment—A Mid-IR and near-IR Spectroscopic Study. Spectrochim. Acta Part A Mol. Biomol. Spectrosc. 2002, 58, 2849–2859. [Google Scholar] [CrossRef]
  67. Jin, B.; Chen, Y.; Pyles, H.; Baer, M.D.; Legg, B.A.; Wang, Z.; Washton, N.M.; Mueller, K.T.; Baker, D.; Schenter, G.K.; et al. Formation, Chemical Evolution and Solidification of the Dense Liquid Phase of Calcium (Bi)Carbonate. Nat. Mater. 2025, 24, 125–132. [Google Scholar] [CrossRef]
  68. Penke, Y.K.; Anantharaman, G.; Ramkumar, J.; Kar, K.K. A Synergistic Cu-Al-Fe Nano Adsorbent for Significant Arsenic Remediation and As(0) Supported Mitigation in Aqueous Systems. In Environmental Arsenic in a Changing World; CRC Press: London, UK, 2019; pp. 457–458. ISBN 978-1-351-04663-3. [Google Scholar]
  69. Puzio, B.; Manecki, M.; Kwaśniak-Kominek, M. Transition from Endothermic to Exothermic Dissolution of Hydroxyapatite Ca5(PO4)3OH–Johnbaumite Ca5(AsO4)3OH Solid Solution Series at Temperatures Ranging from 5 to 65 °C. Minerals 2018, 8, 281. [Google Scholar] [CrossRef]
  70. Li, J.-S.; Wang, L.; Cui, J.-L.; Poon, C.S.; Beiyuan, J.; Tsang, D.C.W.; Li, X.-D. Effects of Low-Alkalinity Binders on Stabilization/Solidification of Geogenic As-Containing Soils: Spectroscopic Investigation and Leaching Tests. Sci. Total Environ. 2018, 631–632, 1486–1494. [Google Scholar] [CrossRef]
  71. Xue, L.; Ren, J.; Wang, S.; Qu, D.; Wei, Z.; Yang, Q.; Li, Y. Preparation of Nanofiber Aerogels by Electrospinning and Studying of Its Adsorption Properties for Heavy-Metal and Dyes. J. Porous Mater. 2020, 27, 1589–1599. [Google Scholar] [CrossRef]
  72. Wang, L.; Cho, D.-W.; Tsang, D.C.W.; Cao, X.; Hou, D.; Shen, Z.; Alessi, D.S.; Ok, Y.S.; Poon, C.S. Green Remediation of As and Pb Contaminated Soil Using Cement-Free Clay-Based Stabilization/Solidification. Environ. Int. 2019, 126, 336–345. [Google Scholar] [CrossRef] [PubMed]
  73. Matsuda, H.; Ishizu, T.; Lee, S.K.; Hasatani, M. Kinetic Study of Ca(OH)2/CaO Reversible Thermochemical Reaction for Thermal Energy Storage by Means of Chemical Reaction. Kagaku Kogaku Ronbunshu 1985, 11, 542–548. [Google Scholar] [CrossRef]
  74. GB 5085.3-2007; Identification Standards for Hazardous Wastes—Identification for Extraction Toxicity. Chinese National Standard: Beijing, China, 2007.
  75. Wang, X.; Zhang, F.; Nong, Z. Mineral Phases and Release Behaviors of As in the Process of Sintering Residues Containing As at High Temperature. Sci. World J. 2014, 2014, 260504. [Google Scholar] [CrossRef] [PubMed]
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Figure 1. (a) Raw soil as collected and (b) the processed soil.
Figure 1. (a) Raw soil as collected and (b) the processed soil.
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Figure 2. Particle size distribution of excavated soil.
Figure 2. Particle size distribution of excavated soil.
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Figure 3. (a) X-ray diffraction (XRD) pattern and (b) TG-DTG analysis of excavated soil.
Figure 3. (a) X-ray diffraction (XRD) pattern and (b) TG-DTG analysis of excavated soil.
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Figure 4. Sample preparation procedures and testing methods.
Figure 4. Sample preparation procedures and testing methods.
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Figure 5. The programmed sintering profile.
Figure 5. The programmed sintering profile.
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Figure 6. Schematic process flow of visualized data after reconstruction.
Figure 6. Schematic process flow of visualized data after reconstruction.
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Figure 7. Compressive test results of sintered excavated soils (representative samples are shown in thumbnails).
Figure 7. Compressive test results of sintered excavated soils (representative samples are shown in thumbnails).
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Figure 8. (a) Arsenic concentration of processed soil, Ca(OH)2 and samples before and after sintering according to Chinese National Standard HJ 803-2016; (b) Leaching characteristics of sintered excavated soil with variable content of Ca(OH)2 in acidic (0.01 M HCl, pH = 2), neutral (ultrapure water, pH = 7) and alkaline (0.01 M NaOH, pH = 12) medium.
Figure 8. (a) Arsenic concentration of processed soil, Ca(OH)2 and samples before and after sintering according to Chinese National Standard HJ 803-2016; (b) Leaching characteristics of sintered excavated soil with variable content of Ca(OH)2 in acidic (0.01 M HCl, pH = 2), neutral (ultrapure water, pH = 7) and alkaline (0.01 M NaOH, pH = 12) medium.
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Figure 9. SEM images: CA0% (a1–a3) before sintering and (b1–b3) after sintering; CA2% (c1–c3) before sintering and (d1–d3) after sintering.
Figure 9. SEM images: CA0% (a1–a3) before sintering and (b1–b3) after sintering; CA2% (c1–c3) before sintering and (d1–d3) after sintering.
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Figure 10. EDS map-scanning of samples before sintering: (a1–a5) CA0%, (b1–b5) CA0.5%, (c1–c5) CA1% and (d1–d5) CA2%; 2-Si Kα1, 3-Al Kα1, 4-Ca Kα1, 5-As Lα1,2.
Figure 10. EDS map-scanning of samples before sintering: (a1–a5) CA0%, (b1–b5) CA0.5%, (c1–c5) CA1% and (d1–d5) CA2%; 2-Si Kα1, 3-Al Kα1, 4-Ca Kα1, 5-As Lα1,2.
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Figure 11. EDS map-scanning of sintered samples: (a1–a5) CA0%, (b1–b5) CA0.5%, (c1–c5) CA1% and (d1–d5) CA2%; 2-Si Kα1, 3-Al Kα1, 4-Ca Kα1, 5-As Lα1,2.
Figure 11. EDS map-scanning of sintered samples: (a1–a5) CA0%, (b1–b5) CA0.5%, (c1–c5) CA1% and (d1–d5) CA2%; 2-Si Kα1, 3-Al Kα1, 4-Ca Kα1, 5-As Lα1,2.
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Figure 12. Compositional ternary diagram of CaO-SiO2-Al2O3 in samples (a) before sintering and (b) after sintering (wt.%).
Figure 12. Compositional ternary diagram of CaO-SiO2-Al2O3 in samples (a) before sintering and (b) after sintering (wt.%).
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Figure 13. Segmented pore phase of samples: (a) Morphology of pores and (b) Arbitrary pore model by equivalent radius (ball size and colors correlate to different equivalent radius).
Figure 13. Segmented pore phase of samples: (a) Morphology of pores and (b) Arbitrary pore model by equivalent radius (ball size and colors correlate to different equivalent radius).
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Figure 14. Pore morphology distributions derived from μCT analysis: (a) Pore equivalent diameter distribution and (b) the corresponding heat map of normalized pore equivalent diameter frequency; (c) Pore volume distribution and (d) the corresponding heat map of normalized volume frequency.
Figure 14. Pore morphology distributions derived from μCT analysis: (a) Pore equivalent diameter distribution and (b) the corresponding heat map of normalized pore equivalent diameter frequency; (c) Pore volume distribution and (d) the corresponding heat map of normalized volume frequency.
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Figure 15. The XRD analysis spectra of samples (a) before and (b) after sintering with different content of calcium hydroxide.
Figure 15. The XRD analysis spectra of samples (a) before and (b) after sintering with different content of calcium hydroxide.
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Figure 16. FTIR spectra of samples (a) before and (b) after sintering with different content of calcium hydroxide.
Figure 16. FTIR spectra of samples (a) before and (b) after sintering with different content of calcium hydroxide.
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Figure 17. As3d XPS spectra of sintered excavated soils: (a) CA0%, (b) CA0.5%, (c) CA1% and (d) CA2%.
Figure 17. As3d XPS spectra of sintered excavated soils: (a) CA0%, (b) CA0.5%, (c) CA1% and (d) CA2%.
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Table 1. Chemical compositions of excavated soil (wt.%).
Table 1. Chemical compositions of excavated soil (wt.%).
SiO2Al2O3K2OFe2O3TiO2MgOSO3P2O5CaO
59.1930.775.202.660.8180.7750.2110.09390.74
Table 2. Sample mixing design.
Table 2. Sample mixing design.
Sample IDSoil (g)Ca(OH)2 (g)Water (g)L/S
CA0%1000300.3
CA0.5%1000.530.150.3
CA1%100130.30.3
CA2%100230.60.3
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Li, X.; Jin, Y.; Wang, Y.; Dong, Z.; Feng, W. Inhibition Mechanism of Calcium Hydroxide on Arsenic Volatilization During Sintering of Contaminated Excavated Soils. Sustainability 2025, 17, 9027. https://doi.org/10.3390/su17209027

AMA Style

Li X, Jin Y, Wang Y, Dong Z, Feng W. Inhibition Mechanism of Calcium Hydroxide on Arsenic Volatilization During Sintering of Contaminated Excavated Soils. Sustainability. 2025; 17(20):9027. https://doi.org/10.3390/su17209027

Chicago/Turabian Style

Li, Xu, Yu Jin, Yaocheng Wang, Zhijun Dong, and Weipeng Feng. 2025. "Inhibition Mechanism of Calcium Hydroxide on Arsenic Volatilization During Sintering of Contaminated Excavated Soils" Sustainability 17, no. 20: 9027. https://doi.org/10.3390/su17209027

APA Style

Li, X., Jin, Y., Wang, Y., Dong, Z., & Feng, W. (2025). Inhibition Mechanism of Calcium Hydroxide on Arsenic Volatilization During Sintering of Contaminated Excavated Soils. Sustainability, 17(20), 9027. https://doi.org/10.3390/su17209027

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