Release and Ecological Risks of Heavy Metals During Coal Combustion in Coal-Fired Power Plants

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This manuscript underscores the necessity for optimized combustion parameters and enhanced particulate filtration systems to mitigate environmental impacts associated with coal-fired power generation.

Abstract

The release of heavy metals during coal combustion may pose potential hazards to the surrounding environment and human health. In this study, we investigated the migration characteristics and ecological risks of heavy metals during the combustion of two distinct raw coal samples (C1 and C2) sourced from Shanxi Province. The analytical results demonstrate significant differences in volatilization behavior between the samples, with total heavy metal release rate ranging from 30.25% to 98.92% for C1 and from 17.77 to 98.16% for C2. Four elements—Cd, As, Pb, and Hg—exhibited preferential migration to fly ash fractions A1 and A2, displaying higher transfer coefficients compared to other monitored heavy metals. Chemical speciation analysis revealed that elemental release behavior was predominantly governed by residual phases (2.2–81.4%), Fe-Mn oxide-bound forms (3.7–45.6%), and sulfate-associated fractions (1.3–56.8%). Combustion temperature showed nonlinear positive correlations with the volatilization rates of Cd, As, Pb, and Hg. Hg volatilization decreases at a combustion temperature below 600 °C, whereas for Cd, As, and Pb, this temperature is below 800 °C. Ecological risk indices (RI) indicate substantial contamination potential in fly ash matrices: A1 (RI = 285.32) is dominated by Hg (I_geo = 1.9, $E_r^i$ = 224) with a notable contribution from Cd ($E_r^i$ = 51), whereas A2 (RI = 246.67) showed a predominance of Cd (I_geo = 1.6, $E_r^i$ = 138) over Hg ($E_r^i$ = 94.4). These findings underscore the need for optimized combustion parameters and enhanced particulate filtration systems to mitigate environmental impacts associated with coal-fired power generation.

1. Introduction

Coal combustion is a primary energy source for many countries, with power plants worldwide burning millions of tons of coal annually [1]. During combustion, heavy metals present in coal are released into the environment through slag, fly ash, and flue gas [2]. Due to their environmental persistence, bioaccumulative properties [3], and inherent toxicity, these heavy metals pose significant risks to both the environment and public health [4,5], thereby hindering progress towards achieving global sustainable development goals [6]. In China, coal remains a dominant component of the national energy mix, with coal-fired power generation accounting for 52.2% of global coal-based electricity production [7]. Extensive research has been conducted on the emission characteristics, speciation patterns, and associated environmental risks of heavy metals during coal combustion. These studies have confirmed that the release of these heavy metals is influenced by multiple factors, including total elemental content, chemical speciation (e.g., organic matter-bound, carbonate-bound, sulfide-bound, and residual phases) [8], and combustion temperature [9,10]. Elements in the mineral-bound form are more strongly influenced by the specific mineral species. The Sequential Extraction Procedure is a multi-step chemical extraction technique [11] that fractionates heavy metals into exchangeable, carbonate-bound, iron-manganese oxide-bound, organic-bound, and residual phases. This method is widely employed to assess the distribution of heavy metals across different chemical forms and has been extensively applied in the analysis of heavy metal speciation in soil, sediment, and wastewater. Wang et al. compared the efficacy of various extraction agents on heavy metal extraction and demonstrated that the sequential chemical extraction method offers higher accuracy in analyzing heavy metal speciation [12]. Tong et al. showed that this method is currently the most widely used approach for investigating the chemical fractionation of heavy metals in coal-fired boiler waste [6]. Zhou et al. conducted a study on low-rank coal from Xinjiang Province, revealing that various heavy metal species—including arsenic (As), lead (Pb), mercury (Hg), and cadmium (Cd)—influence the release behavior of these metals, with distinct emission pathways observed under high-temperature conditions [8]. For instance, As, Pb, and Hg exhibit a high affinity for sulfides [8], whereas Cd predominantly occurs in organically bound and residual forms in coal [13]. Heavy metals bound to organic matter or sulfides volatilize more readily during combustion than their mineral-bound counterparts, whose release is influenced by the mineral composition [14]. The release of heavy metals during combustion is significantly influenced by process temperature. At lower temperatures (<400 °C), organically bound metals tend to be released [14,15], as demonstrated by studies on waste incineration and incineration fly ash. As temperatures rise, release rates rise nonlinearly. For instance, Hg volatilizes rapidly at 600 °C due to its high volatility [16]. Carbonate- and sulfide-bound metals are substantially emitted below 800 °C, whereas those associated with clay minerals require higher temperatures (>1000 °C) for gradual release [13,14,15].

Studies by Lu et al. have shown that increasing temperature can promote the release of heavy metals, with high temperatures of 1000–1100 °C having a significantly stronger impact on heavy metal release than low temperatures of 700–900 °C [17]. Furthermore, different combustion atmospheres (O₂/CO₂) significantly influence the release behavior of heavy metals. Studies have shown that a CO₂-rich atmosphere can reduce the volatility of heavy metals and suppress their emission into the environment [17]. Zhao et al. investigated the release characteristics of alkali metal salts under O₂/CO₂ combustion conditions using FES technology and demonstrated that the CO₂ atmosphere promotes the conversion of NaAc, NaCl, and similar compounds into Na₂CO₃, which exhibits lower volatility and further inhibits the release of gaseous sodium (Na(g)) [18].

Heavy metals such as Hg, As, Pb, and chromium (Cr) emitted during coal combustion have been shown to exert significant ecological impacts, as demonstrated by studies on the emission and control of these pollutants. Fan et al. [19] reported that Cd in soils near coal-fired plants predominantly occurs in acid-extractable and reducible fractions, posing acute environmental risks. Studies have revealed severe Cd contamination in the topsoils surrounding coal-fired power plants in southern China, with Cd concentration significantly influenced by the direction of the prevailing winds [20]. Gypsum exhibits higher environmental risks than fly ash with respect to Hg, Cr, Cd, Pb, and As, owing to its lower environmental stability [5,21]. Sun et al. [22] confirmed coal combustion as the dominant source of heavy metals emissions during both heating and non-heating seasons, particularly in Chifeng City, northern China [23]. Coal-derived Hg, characterized by high volatility, low aqueous solubility, and atmospheric persistence of up to two years, contributes significantly to global Hg pollution through long-range atmospheric transport [24]. According to the United Nations Environment Programme, coal combustion is one of the primary sources of anthropogenic Hg emissions, accounting for approximately 25% of total global emissions, with the coal-fired power and heating sectors representing the largest anthropogenic Hg emission sources [25].

Given the long-term ecological risks associated with cumulative heavy metal release, understanding the emission mechanisms and potential environmental impacts of these pollutants remains a critical research priority.

Previous studies have predominantly focused on laboratory-simulated combustion processes, leaving a critical knowledge gap in the characterization of actual emissions and associated ecological risks from operational coal-fired power plants.

(a)

Shanxi Province, recognized as China’s primary coal production base, faces escalating environmental challenges arising from coal combustion activities.

(b)

This investigation examines two representative coal-fired power plants in Shanxi using an integrated methodology combining field emission sampling, advanced laboratory analysis, and systematically controlled combustion experiments. Enrichment factors, geo-accumulation indices, long-range migration potential, and ecological risk assessments were used to address the following three research questions: (1) What is the migration rate of heavy metals in coal during the combustion process, from raw coal to slag and fly ash? (2) How do the chemical forms of heavy metals in raw coal and combustion temperature influence their migration behavior? (3) Considering the content, speciation, and leaching characteristics of heavy metals in slag and fly ash, what is their potential ecological risk level?

(c)

The findings of this study not only provide predictive insights into possible ecological risks associated with local coal combustion but also serve as a theoretical foundation for developing targeted management strategies to mitigate heavy metal pollution. Furthermore, the results are applicable to predicting heavy metal release during coal combustion in regions with high coal production and similar coal properties.

2. Materials and Methods

2.1. Study Area

Power Plant 1 is situated in Jinzhong City, Shanxi Province, and utilizes lean coal sourced from Carboniferous–Permian coal-bearing strata within the Xishan Coalfield. Power Plant 2 is located in the Datong Coalfield of Shanxi Province and employs gas coal extracted from analogous Carboniferous–Permian geological formations. Both coal types are representative of medium metamorphic bituminous coal. Power Plant 1 operates at a rated capacity of 1000 t·h⁻¹ and employs an external flue gas treatment system to control emissions prior to atmospheric release. In contrast, Power Plant 2 operates at a rated capacity of 700 t·h⁻¹ and features an integrated indoor emission control system with internally positioned flue gas outlets. Both facilities are equipped with pulverized coal-fired boilers and maintained operational loads between 85% and 90% during the sampling period.

2.2. Sampling and Chemical Analysis

2.2.1. Sampling

Raw coal (C1, C2), slag (S1, S2), and fly ash (A1, A2) samples, each weighing approximately 200 g, were collected from the two power plants. All samples were air-dried to constant weight under ambient conditions (25–28 °C). Subsequently, raw coal and slag samples were pulverized to a particle size of 200 mesh for analytical purposes.

2.2.2. Chemical Analysis

The industrial analysis of coal was conducted in accordance with the national standard of the People’s Republic of China, GB/T 212-2008, the “Method for Industrial Analysis of Coal” [26].

Approximately 0.25 g of homogenized sample powder was accurately weighed into a specialized graphite digestion vessel (Horde, HD-SM20, Weifang, China). A 5 mL aliquot of aqua regia solution (prepared with a HCl-to-HNO₃ volume ratio of 4:1) was then added for acid digestion. The mixture was thoroughly homogenized and subjected to a 2-h pre-digestion period at ambient temperature. Complete digestion was subsequently achieved through controlled heating using a graphite-block digestion system. After digestion, the solution was diluted with ultrapure water to achieve a final acid matrix concentration of 2% (v/v). Quantitative analysis of thirteen heavy metal elements—aluminum (Al), iron (Fe), manganese (Mn), zinc (Zn), copper (Cu), titanium (Ti), Cd, Cr, nickel (Ni), Pb, As, selenium (Se), Hg—was performed using high-resolution inductively coupled plasma mass spectrometry (ICP-MS, PerkinElmer NexION 300X, Beijing, China), with instrumental conditions optimized for multi-element detection. The standard addition method was employed to ensure analytical accuracy and precision.

2.3. Calculation of Heavy Metal Release Rates

The release rates of heavy metals to slag (T_slag) and fly ash (T_ash) were calculated according to the following equation [27]:

$$T_{slag}={\displaystyle \frac{C_{slag}\times m_{slag}}{C_{coal}\times m_{coal}}}\times 100\%$$

(1)

$$T_{ash}={\displaystyle \frac{C_{ash}\times m_{ash}}{C_{coal}\times m_{coal}}}\times 100\%$$

(2)

where C_slag, C_ash, and C_coal denote the concentrations of heavy metal in slag, fly ash, and raw coal, respectively.

2.4. Speciation Analysis of Heavy Metals

Four trace elements (Cd, As, Pb, Hg) were systematically selected for sequential chemical extraction (SCE) analysis using the standardized Tessier method [28,29]. Representative samples, including raw coal, slag, and fly ash, were subjected to the sequential analytical procedures outlined below (all reagents used in this study were purchased from Tianjin Keminuo Chemical Reagent Co., Ltd., located in Tianjin, China.):

Acid-soluble fraction (F1): 1 g of sample was mixed with 30 mL of 1 mol·L⁻¹ CH₃COONH₄ (pH adjusted to 5 using CH₃COOH). The mixture was shaken at 120 rpm for 16 h, centrifuged at 2000 rpm for 21 min, and the supernatant was collected. The residue (R1) was washed, dried at 50 °C, and retained for further analysis.

Reducible fraction (F2): The residue R1 was treated with 30 mL of 0.1 mol·L⁻¹ NH₂OH·HCl in 25% HNO₃. The mixture was shaken for 16 h at 120 rpm, centrifuged at 2000 rpm for 21 min, and the supernatant was collected. The residue (R2) was dried at 50 °C and retained for further analysis.

Oxidizable fraction (F3): The residue R2 was mixed with 10 mL of ultrapure water (pH adjusted to 2 using HNO₃) and 10 mL of 30% H₂O₂. The mixture was heated at 85 °C for 1 h, evaporated to near dryness, then treated with 30 mL of 1 mol·L⁻¹ CH₃COONH₄. This suspension was shaken at 120 rpm for 16 h and centrifuged at 2000 rpm for 21 min. The supernatant was then collected. The residue (R3) was dried at 50 °C and retained for further analysis.

Residual fraction (F4): The residue R3 was ashed at 500 °C, followed by microwave-assisted digestion using a HNO₃: HF mixture with a volume ratio of 5:2. The resulting solution was then analyzed using inductively coupled plasma mass spectrometry (ICP-MS).

Blank samples were included in each procedural step to monitor potential contamination. All extracts were stored at 4 °C prior to analysis.

2.5. Combustion Temperature Gradient Experiment

2.5.1. Experimental Procedure

Pulverized C1 and C2 samples (1 g each) were combusted in a muffle furnace at temperatures of 400 °C, 600 °C, 800 °C, 1000 °C, and 1200 °C, with a heating rate of 10 °C/min and isothermal holding at each temperature for 2 h. The heating rate applied in the experiment (10 °C/min) corresponds to the typical range of temperature increase of 20–25 °C·h⁻¹ observed in industrial boilers. Post-combustion residues were cooled to ambient temperature, weighed accurately, and ground to a particle size of 200 mesh (74 µm) for subsequent analytical characterization.

2.5.2. Analytical Method

Slag samples were subjected to acid digestion, and the concentrations of Cd, As, Pb, and Hg were determined using inductively coupled plasma mass spectrometry (ICP-MS). The proportion of each metal released to flue gas (D) was calculated as follows [30]:

$$D={\displaystyle \frac{(c_1\times m_1-c_2\times m_2)}{c_1\times m_1}}\times 100\%$$

(3)

where c₁ and m₁ denote the heavy metal concentration in raw coal (mg·kg⁻¹) and the sample mass (1 g), respectively, and c₂ and m₂ represent the concentration and mass of the metal in slag.

2.6. Assessment Indices

2.6.1. Enrichment Factor (EF)

The enrichment factor (EF) is a widely used indicator for assessing the extent of anthropogenic pollution by comparing the concentrations of heavy metal elements in soil with their respective geochemical background values. To minimize inter-sample variability, EF calculations employ a reference element with low natural variation. In this study, Al was selected as the reference element to distinguish between combustion-induced enrichment and natural geogenic contributions. The rationale for choosing Al is twofold: (1) it has relatively high abundance in the Earth’s crust and is ubiquitously distributed, thus representing the baseline crustal composition; and (2) it exhibits stable chemical properties and is minimally influenced by anthropogenic activities. The EF quantifies the degree of heavy metal enrichment in fly ash relative to average crustal abundances [31]. However, it should be noted that the formula provided in this context does not correspond to the standard EF calculation; the correct expression is given in Equation (4): [31]

$$\mathrm{E}\mathrm{F}={\displaystyle \frac{C_{ash}/{Al}_{ash}}{C_{coal}/{Al}_{coal}}}$$

(4)

where C_ash denotes the concentration of heavy metals in fly ash (mg·kg⁻¹); Al_ash represents the Al concentration in fly ash (mg·kg⁻¹); C_coal refers to the concentration of heavy metals in raw coal (mg·kg⁻¹); and Al_coal indicates the Al concentration in raw coal (mg·kg⁻¹).

2.6.2. Geoaccumulation Index (I_geo)

The geoaccumulation index (I_geo) is used to quantify the extent of pollutant accumulation in a specific region and to assess associated environmental exposure risks [32,33]. This method takes into account both anthropogenic contamination and environmental geochemical influences on background values, as well as natural geogenic contributions to baseline concentrations.

The I_geo is calculated according to the following equation [32]:

$$I_{geo}={log}_2\left({\displaystyle \frac{C_n}{1.5\times B_n}}\right)$$

(5)

where C_n denotes the measured concentration of the heavy metal (mg·kg⁻¹), B_n represents the regional geochemical background value specific to Shanxi Province, and a correction factor of 1.5 is applied to account for natural variability in these background values.

The classification of pollution severity is as follows: I_geo ≤ 0, unpolluted; 0 < I_geo ≤ 1, mild pollution; 1 < I_geo ≤ 2, moderate pollution; 2 < I_geo ≤ 3, moderately severe pollution; 3 < I_geo ≤ 4, severe pollution; 4 < I_geo ≤ 5, extreme pollution; I_geo > 5, catastrophic pollution.

2.6.3. Long-Range Atmospheric Transport

The Gaussian dispersion model was systematically applied to simulate the long-range atmospheric transport dynamics of heavy metal contaminants emitted during prolonged coal combustion, enabling a comprehensive assessment of potential ecological risks and their spatiotemporal distribution patterns. Key assumptions included: coal combustion rate (m_coal), 1000 kg·s⁻¹; wind speed (u), 2 m·s⁻¹ (ranging from 1.9 to 2.1 m·s⁻¹ in the study area); effective emission height (H), 50 m; and atmospheric stability under neutral class, with dispersion parameters defined as ${\sigma }_y=0.2x^{0.9}, {\sigma }_z=0.1x^{0.9}$.

The Gaussian model formula (Equation (6)) is expressed as follows [34]:

$$C_{\left(x,y,z\right)}={\displaystyle \frac{Q}{2\pi u{\sigma }_y{\sigma }_z}}\mathit{exp}\left(-{\displaystyle \frac{y^2}{2{\sigma }_y^2}}\right)\left[\mathit{exp}\left(-{\displaystyle \frac{{\left(z-H\right)}^2}{2{\sigma }_z^2}}\right)+\mathit{exp}\left(-{\displaystyle \frac{{\left(z+H\right)}^2}{2{\sigma }_z^2}}\right)\right]$$

(6)

where C(x,y,z) denotes the heavy metal concentration at spatial coordinates (mg·m⁻³); Q represent the source strength (mg·s⁻¹), calculated as $Q=C_{coal}\times (1-ƞ)\times f\times m_{coal}$; u is the wind speed (m·s⁻¹); σ_y and σ_z are the horizontal and vertical diffusion coefficients, respectively, which depend on atmospheric stability; H is the effective emission height (m); C_coal refers to the native heavy metal concentration in coal (mg·kg⁻¹); ƞ denotes the combustion efficiency (90%); and f represents the volatility factor. Non-volatile elements (Al, Si, Fe, Ti) were excluded from long-range transport analysis. Volatility factors for other metals were assigned according to the following reference [34]:

f_Mn = 0.05, f_Zn = 0.45, f_Cu = 0.2, f_Cd = 0.7, f_Cr = 0.1, f_Ni = 0.1, f_Pb = 0.6, f_As = 0.8, f_Se = 0.9, f_Hg = 0.95.

2.6.4. Potential Ecological Risk Index (RI)

The potential ecological risk index (RI), originally developed by Hakanson (1980) [11], not only considers heavy metal concentrations but also integrates ecological, environmental, and toxicological impacts, employing a standardized and unified attribute index classification system. This methodology synthesizes four critical parameters: metal toxicity coefficients, environmental concentration gradients, synergistic interactions among contaminants, and ecosystem sensitivity to pollutant exposure. [35].

The single-element ecological risk factor ($E_r^i$) and the comprehensive risk index (RI) are calculated as follows [35]:

$$E_r^i=T_r^i\times C_f^i$$

(7)

$$RI={\sum }_{i=1}^n{E}_r^i$$

(8)

where $E_r^i$ denotes the potential ecological risk factor for the heavy metal; $T_r^i$ represents the toxicity response coefficient of heavy metal i, reflecting its inherent toxicity; $C_f^i$ is the contamination factor of heavy metal i, calculated as C_sample/B_n (where B_n is the regional background value); and n refers to the number of heavy metals evaluated. The toxicity response coefficients ($T_r^i$) were assigned as follows: Zn = 1; Cu = Ni = Pb = 5; Cd = 30; Cr = 2; As = 10; Hg = 40; Hg = 40 [11].

2.7. Statistical Analysis

Data processing and graphical plotting were performed using SigmaPlot 13.0 software (Systat Software, Inc., San Jose, CA, USA), and differential analysis was conducted through a one-way ANOVA. Significant differences were determined by a Tukey HSD test with a level of significance of 95% (p < 0.05). One-way ANOVA was conducted to compare the migration rates of 13 heavy metals from samples C1 and C2 to slag and fly ash. Similarly, the leaching rates of Cd, As, Pb, and Hg in samples C1, C2, S1, S2, A1, and A2 during the indoor combustion test were also analyzed using one-way ANOVA. The volatilization rates of Cd, As, Pb, and Hg in flue gas under different combustion temperatures for samples A1 and A2 were also evaluated by one-way ANOVA. All analyses were performed using SPSS 20.0 (IBM, Armonk, NY, USA).

3. Results and Discussion

3.1. Key Coal Quality Parameters of the Raw Coal Samples

Key coal quality parameters, including moisture content, ash content, volatile matter, and calorific value, are presented in

Table 1. Key coal quality parameters of the raw coal samples.

Samples	Proximate Analysis				Ultimate Analysis
	Mad	Ad	St,d	Vdaf	Carbon	Hydrogen	Nitrogen	Oxygen
	(%)	(%)	(%)	(%)	(%)	(%)	(%)	(%)
C1	1.01	22.09	2.54	14.78	57.8	2.7	1.5	4.8
C2	2.77	21	1.01	39.71	55.9	3.1	1.4	2.2

M_ad, moisture (air-dried basis); A_d, ash content (dry basis); S_t,d, total sulfur (dry basis); V_daf, volatile matter (dry ash-free basis).

.

3.2. Migration of Heavy Metal Elements After Raw Coal Combustion

The concentrations of heavy metal elements in raw coal (C1 and C2), slag (S1 and S2), and fly ash (A1 and A2) from two coal-fired power plants were determined using ICP-MS. The migration rates of heavy metals from raw coal to slag and fly ash following combustion were calculated according to Equations (1) and (2). As shown in the results (Figure 1), the total release rates of heavy metals in C1 ranged from 30.25% to 98.92%, while those in C2 varied between 17.77% and 98.16%, indicating that environmental conditions significantly influence release dynamics. For the thirteen heavy metals analyzed, the proportions of migration from C1 to slag (S1) and fly ash (A1) were 0.92–47.38% and 21.79–78.22%. Correspondingly, the migration proportions from C2 to slag (S2) and fly ash (A2) ranged from 2.04 to 47.43% and from 12.99 to 90.98%, respectively.

Comparative analysis revealed statistically significant increases in the total release rates of Cr, Ni, Se, and Hg from C1 compared to C2 (p < 0.05). In contrast, Mn, Cu, Ti, Cd, Pb, and As exhibited higher cumulative release from C2 (p < 0.05). Given that Al, Fe, Mn, and Ti were present at high concentrations in raw coal, slag, and fly ash, significant linear correlations were observed between C1 and A1/S1, as well as between C2 and A2/S2 (Figure S1). This suggests that elevated concentrations of Al, Fe, Mn, and Ti in C1 and C2 result in a proportional increase in their content in the corresponding slag and fly ash. The correlation analysis results for the remaining nine elements are summarized in Figure S2: a linear correlation was observed between the heavy metal concentrations in C1 and A1, whereas a nonlinear correlation was identified between C1 and S1, consistent with findings from previous studies. In the C2 coal seam, significant nonlinear correlation relationships were found between raw coal and the indicators A2 and S2, as determined by statistical analysis of coal quality data using standard software. These results indicate that increasing heavy metal concentrations in raw coal enhanced their migration to both slag and fly ash in C1, whereas in C2, higher concentrations primarily promoted migration to slag.

Notably, Power Plant 2 employs an internal flue gas recirculation system, which may allow a portion of the fly ash to settle into the slag, thereby complicating the interpretation of heavy metal release mechanisms in C2. Both power plants exhibited low capture efficiencies for Hg, likely due to its high volatility, making the capture of elemental Hg during combustion challenging and resulting in substantial gaseous Hg emissions. However, the Hg concentrations measured in the slag from Plant 2 were significantly higher than those from Plant 1, which is possibly attributable to the design of its internal flue gas recirculation system. Tang et al. investigated the migration behavior of trace elements (Be, V, Cr, Co, Ni, Cu, Zn, As, Se, Mo, Cd) from coal to gasification residues using a five-step sequential chemical extraction procedure. They highlighted that the differing forms of heavy metals may result from distinct transformation pathways during the combustion process, which thereby influence their release behavior [36].

Previous studies have consistently identified Cd, As, and Pb as key indicators for coal combustion [37]. Our findings are consistent with these, demonstrating significant migration rates of Cd, As, and Pb to fly ash in both coal types, with particularly pronounced migration from C2 to A2 [37]. Based on actual coal utilization in thermal power plants, the raw coal samples employed in this study were derived from a composite mixture of materials sourced from multiple coal seams. The analysis of these blended coal samples not only conforms to real-world engineering practices but also accurately represents the release characteristics of heavy metals during the combustion of a defined coal quantity.

3.3. Speciation of Heavy Metals in Raw Coal, Slag, and Fly Ash

Studies have demonstrated that the release of heavy metals during coal combustion is closely linked to their chemical speciation. To evaluate the environmental impact of heavy metals in coal combustion byproducts, sequential chemical extraction (SCE) was employed to analyze the speciation of four heavy metals—Cd, As, Pb, and Hg—in raw coal, slag, and fly ash collected from two coal-fired power plants. The extraction protocol targeted four distinct fractions: F1 (acid-soluble fraction), encompassing the carbonate-bound phase in raw coal and exchangeable phase in slag/fly ash; F2 (reducible fraction), representing the sulfide-bound phase in raw coal and Fe/Mn oxide-bound phase in slag/fly ash; F3 (oxidizable fraction), corresponding to the organic matter-bound phase in raw coal and sulfate-bound phase in slag/fly ash; and F4 (residual fraction), primarily composed of silicate minerals dominated by clay minerals [11].

The recovery rates of heavy metals using the four-step sequential extraction method ranged from 85% to 119%, consistent with established standards for trace-level analysis. The concentration levels of four chemical elements in raw coal, slag, and fly ash, as determined by the continuous leaching method, are as follows: (1) Cd concentrations: 0.407 mg·kg⁻¹ and 0.543 mg·kg⁻¹ in C1 and C2, respectively; 0.0538 mg·kg⁻¹ and 0.03 mg·kg⁻¹ in S1 and S2, respectively; and 0.15 mg·kg⁻¹ and 0.344 mg·kg⁻¹ in A1 and A2, respectively.

(2) As concentrations: 4.81 mg·kg⁻¹ and 1.85 mg·kg⁻¹ in C1 and C2, respectively; 0.805 mg·kg⁻¹ and 0.0743 mg·kg⁻¹ in S1 and S2, respectively; and 3.78 mg·kg⁻¹ and 1.38 mg·kg⁻¹ in A1 and A2, respectively.

(3) Pb concentrations: 18.2 mg·kg⁻¹ and 22.7 mg·kg⁻¹ in C1 and C2, respectively; 4.586 mg·kg⁻¹ and 1.52 mg·kg⁻¹ in S1 and S2, respectively; and 9.043 mg·kg⁻¹ and 20.456 mg·kg⁻¹ in A1 and A2, respectively.

(4) Hg concentrations: 0.426 mg·kg⁻¹ and 0.476 mg·kg⁻¹ in C1 and C2, respectively; 0.005 mg·kg⁻¹ and 0.02 mg·kg⁻¹ in S1 and S2, respectively; and 0.2 mg·kg⁻¹ and 0.065 mg·kg⁻¹ in A1 and A2, respectively.

A strong linear correlation was observed between the sum of sequentially extracted metal fractions and the total metal concentrations determined by ICP-MS (y = 1.04x + 0.04, R² = 0.99) (Figure S3), confirming the method’s analytical reliability. The speciation profiles of Cd, As, Pb, and Hg in raw coal (C1, C2), slag (S1, S2), and fly ash (A1, A2), obtained through sequential chemical extraction, are presented in Figure 2.

3.3.1. Speciation of Cd

In both raw coal samples (C1 and C2), Cd was predominantly associated with the organic matter-bound phase (43.3% and 48.5% of the extracted totals, respectively) and the residual fraction (35.4% and 30.4%, respectively), with smaller proportions present in the sulfate-bound phase (16.6% and 17.2%, respectively). Notably, there was no significant difference in the distribution of Cd species between C1 and C2. During the transformation of Cd from coal to slag, the carbonate-bound, sulfide-bound, and organic matter-bound phases were largely converted into the residual fraction, accounting for 59.4% in S1 and 56.6% in S2, followed by the Fe/Mn oxide-bound phase (30.5% in S1 and 34.7% in S2). In contrast, Cd in fly ash exhibited a more diversified speciation pattern, with dominant contributions from the Fe/Mn oxide-bound phase (24.7% in A1 and 22.3% in A2), the sulfate-bound phase (33.4% in A1 and 34.6% in A2), and the residual fraction.

3.3.2. Speciation of As

In both unprocessed coal specimens (C1 and C2), As was predominantly associated with the sulfide-bound phase (42.5% and 38.7%, respectively) and residual fraction (45.6% and 48.5%, respectively), following the order: residual fraction > sulfide-bound phase > organic matter-bound phase. Significant differences were observed in the proportions of the carbonate-bound and organic matter-bound phases between C1 and C2, whereas no significant differences were detected in the sulfate-bound phase or the residual fraction. During the transformation of As from coal to slag and fly ash, the majority of As was converted into the residual fraction (79.4% in S1, 81.4% in S2, 72.2% in A1, and 78.0% in A2). The limited presence of the sulfate-bound phase in fly ash may be attributed to the preferential formation of arsenates by reactions between As oxides and heavy metal oxides under high-temperature conditions [38].

3.3.3. Speciation of Pb

Pb was distributed across all four fractions in both coal types, with the carbonate-bound phase being one of the dominant forms. Moreover, no significant differences were observed in the distribution of Pb species between C1 and C2. During the transformation of Pb from coal to slag, the predominant phases were the exchangeable phase, constituting 50.2% in S1 and 48.9% in S2, and the Fe/Mn oxide-bound phase, accounting for 40.2% in S1 and 44.2% in S2, while the residual fraction remained relatively small. This pattern likely reflects the inherently low content of residual Pb in raw coal and the thermal instability of other Pb compounds at high temperatures, resulting in near-complete volatilization. Additionally, slag tends to enrich Fe/Mn hydroxides, which further inhibits the formation of residual Pb during combustion. In high-temperature flue gas zones, Pb predominantly migrated to solid particles through sulfate-mediated condensation due to its strong affinity for sulfate-induced solidification [39]. In fly ash, Pb was primarily present in the residual fraction, representing 66.9% in A1 and 70.4% in A2, with smaller proportions associated with the Fe/Mn oxide-bound and sulfate-bound phases. The scarcity of exchangeable Pb is attributed to gaseous Pb species reacting with aluminosilicate minerals in fly ash under elevated temperatures, leading to chemical immobilization [33].

3.3.4. Speciation of Hg

In both coal samples, Hg was predominantly associated with the sulfide-bound phase (47.22% in C1 and 40.7% in C2) and the organic matter-bound phase (39.5% in C1 and 32.4% in C2), consistent with its known affinity for pyritic minerals in bituminous coal [7,40], which commonly host Hg in sulfide phases [40]. Notably, significant differences were observed in the proportions of the carbonate-bound and organic matter-bound phases between C1 and C2, whereas no significant differences were detected in the sulfate-bound phase or the residual fraction. During the transformation of Hg from coal to slag, the Fe/Mn oxide-bound phase (45.6% in S1 and 37.6% in S2) and the residual fraction (40.6% in S1 and 50.2% in S2) were dominant, with minimal contribution from the sulfate-bound phases. In fly ash, Hg was primarily present in the sulfate-bound (56.8% in A1 and 47.5% in A2) and residual fractions (37.9% in A1 and 45.6% in A2).

The speciation patterns of four heavy metal elements (Cd, As, Pb, Hg) in raw coal samples are largely consistent with those reported in previous studies. As, Pb, and Hg exhibit strong affinities for sulfide-bound phases, whereas Cd preferentially partitions into organic matter-bound and residual fractions. Sequential chemical extraction results confirm that As and Pb in raw coal are predominantly associated with sulfides, while Hg mainly occurs in sulfide-bound and residual forms. Notably, heavy metal speciation in coal exhibits significant regional variability, primarily influenced by differences in coal-forming materials and coalification processes [3]. Nevertheless, the two bituminous coal samples examined in this study display comparable speciation characteristics, a finding likely attributable to their similar coalification histories.

Both coal-fired power plants utilized pulverized coal boilers operating at 1200 °C, a temperature sufficient to volatilize nearly all heavy metal phases. The post-combustion speciation of Cd, As, Pb, and Hg in slag and fly ash was strongly influenced by combustion temperature, atmosphere, coal particle fineness, and initial forms present in raw coal. Despite differences in absolute concentrations between the two coal samples (Figure 2), combustion under identical conditions (1200 °C) resulted in similar transformation patterns for all four metals. Further analysis revealed that organically bound phases exhibited significantly higher volatility than mineral-bound phases under high-temperature combustion, with the latter’s volatilization governed by the specific mineralogical composition of the associated phases [3].

The environmental mobility and toxicity of heavy metals in slag and fly ash are determined by their chemical speciation. Labile fractions, such as the exchangeable and acid-soluble phases, significantly enhance metal mobility and thereby increase the potential for contamination of surrounding ecosystems. In contrast, the residual fraction is characterized by low bioavailability and minimal environmental impact.

3.4. Migration of Four Heavy Metals Under Varied Combustion Temperatures

The release of heavy metals during coal combustion is influenced not only by their chemical speciation in coal but also by combustion temperature [41]. According to relevant reports [42], the primary chemical forms of the 13 heavy metals in raw coal are as follows: aluminum silicate and Al₂O₃; FeS₂ and Fe₂O₃; MnO₂, MnCO₃, MnS, and MnCl₂; ZnS, ZnO, and Zn²⁺ [42,43]; CuS and Cu; TiO₂; CdS and CdCO₃; organic-bound Cr, Fe-Mn-Cr mixed oxides, and Cr₂S₃ [34,42]; NiS, NiO, and Ni(OH)₂; PbCO₃ and PbS; As–S compounds, Ca₃(AsO₄)₂, FeAsO₄, and AlAsO₄ [14]; SeO₂ [14]; and Hg, HgS, and HgCl₂. Following coal combustion, the predominant species of these heavy metals are observed to be: Al₂O₃ (s) and Al³⁺ (l) [44]; Fe₂O₃ and Fe₃O₄ (s) [45]; MnO₂ and Fe-Mn oxides (s) [33]; Zn(g), ZnCl₂(g), ZnO, and ZnS (s) [43]; CuO(s), Cu₂O(s), and CuCl₂(g) [46]; TiO₂(s) and Ti⁴⁺ (l); Cd(g), CdO(s), and CdS(s) [47]; Cr₂O₃ and CrO₃(g) [47]; NiO(s) and NiCl₂(g) [46]; Pb(g), PbCl₂(g), PbO(s), and PbS(s) [14,47]; As₂O₃(g), As₂O₅(g), and As₂O₅(s) [48]; Se(g) and SeO₂(g) [14]; and Hg⁰(g), HgCl₂(g), HgO(s), and HgS(s) [42].

During combustion, Cd, As, Pb, and Hg are volatilized into flue gas as a result of thermal decomposition. The volatilization rates of these metals from raw coals C1 and C2 under varying combustion temperatures were calculated using Equation (3), with results illustrated in Figure 3. Previous studies have demonstrated that during coal pyrolysis, elements such as Cd, As, and Hg exhibit elevated volatilization at specific temperature ranges. Notably, no significant differences were observed in the volatilization behavior of the four metals between C1 and C2. The volatilization rates of Cd, As, and Pb increase nonlinearly with rising combustion temperature, consistent with findings from previous studies on their emission characteristics. Hg volatilization begins to decline below 600 °C, whereas the corresponding thresholds for Cd, As, and Pb occur below 800 °C.

Combustion temperature significantly influences the release dynamics of heavy metals during thermal processes. Elevated temperatures promote chemical interactions between hydrogen chloride (HCl) in flue gas and metallic species, leading to the formation of volatile chloride compounds that enhance metal mobilization [49]. When temperatures exceed 1000 °C, Cd, As, and Pb exhibit reactive affinity toward aluminosilicate matrices in fly ash, resulting in the formation of thermally stable compounds such as cadmium sulfide (CdS), cadmium oxide (CdO), arsenic pentoxide (As₂O₅), and lead sulfate (PbSO₄), thereby inhibiting further volatilization. This dual mechanism—comprising chloride-induced volatilization temperature phenomena stabilization—gives rise to a distinct, phase-specific evolution pattern in heavy metal speciation within flue gas systems.

3.4.1. Release Behavior of Cd

Cd occurs in coal in both organic and inorganic forms. At temperatures below 400 °C, organically bound Cd undergoes volatilization. Between 400 °C and 800 °C, carbonate-bound and organic-bound Cd are released from raw coal, with volatility increasing proportionally with temperature. In contrast, Cd associated with sulfide minerals and clay minerals requires higher temperatures (approximately 1000 °C) for gradual release. In flue gas, Cd predominantly exists as elemental Cd, CdO, Cd(OH)_x, and CdS [34].

3.4.2. Release Behavior of As

As occurs in coal in both organic and inorganic forms, and is predominantly present in the latter as solid As₂O₃. Organic As is released during combustion through oxygen- and sulfur—mediated transformation processes. Below 400 °C, organic As undergoes volatilization [34]. Between 400 °C and 600 °C, As₂O₃ (g) is the dominant species in flue gas, accompanied by partial oxidation to As₂O₅(g) [50]. Between 600 °C and 800 °C, As₂O₃(g) remains prevalent, while a portion is converted to AsCl₃(g) via chlorination reactions. Between 800 °C and 900 °C, multiple species coexist, including AsO(s), AsH₃, As₄O₆(s), and As₂O₅(g) [8]. Above 900 °C, AsO(s) becomes the dominant phase, whereas temperatures exceeding 1300 °C favor the stabilization of As₂O₃ and As₂O₅ [50]. The speciation of As in flue gas includes AsO, AsH₃, As₄O₆(s), As₂O₅(g), and AsCl₃(g).

3.4.3. Release Behavior of Pb

Pb occurs in coal in both organic and inorganic associations. Below 400 °C, Pb demonstrates negligible volatilization (<10%). In the intermediate temperature range of 400 °C to 800 °C, sulfide-bound and carbonate-bound Pb undergoes chlorination reactions with HCl or Cl₂ to form volatile PbCl₂ [10], leading to increased volatilization rates that rise proportionally with temperature. At temperatures exceeding 800 °C, aluminosilicate-associated Pb undergoes progressive thermal release. The speciation of Pb in flue gas is dominated by PbCl, elemental Pb, PbO, PbH, PbCl₂, and PbSO₄ species [8]. Combustion temperature governs Pb speciation as follows: between 400 °C and 600 °C, PbO(s) is the dominant species, with minor contributions from PbCl₂ and PbSO₄ [51]; from 600 °C to 1200 °C, PbO(s) vaporizes to form PbO(g); above 1200 °C, partial decomposition of PbO(g) yields Pb(g), resulting in a coexisting mixture of PbO(g) and Pb(g) [51,52].

3.4.4. Release Behavior of Hg

Hg occurs in coal in both organic and inorganic forms. Due to the low boiling point of organic Hg compounds, a substantial portion is released during the thermal decomposition of organic matter at temperatures below 300 °C. By 400 °C, approximately 50% of Hg has been volatilized [9]. Strezov et al. [53] observed complete release of organic-bound Hg at 400 °C, where it reacts with volatile species such as H₂, HCl, and H₂S in volatile products to form HgCl₂ [9]. Between 400 °C and 600 °C, sulfide-bound Hg undergoes decomposition, accompanied by a decrease in HgCl₂ concentrations. Above 800 °C, nearly all Hg is converted to elemental Hg (Hg⁰), which remains stable in flue gas [54]. Studies indicate that at a combustion temperature of 600 °C, over 90% of Hg is released from raw coal into the flue gas, with volatilization rates increasing as temperature rises.

In summary, when temperatures fall below 400 °C, heavy metals associated with organic matter are released, particularly those with a strong affinity for organics, such as Cd and Pb, which exhibit significant volatilization. By 600 °C, over 90% of Hg has been volatilized, while at 800 °C, most organically bound metals are emitted, leaving the residual fractions as the dominant phase in the solid residue. Above 1000 °C, residual fractions undergo gradual release, with approximately 90% of Cd, 85% of Hg, and 80% of Pb ultimately migrating from coal to flue gas.

The post-combustion speciation of Cd, As, Pb, and Hg in flue gas directly determines their environmental impact. For example, volatile species such as CdCl₂ and Hg⁰ pose higher atmospheric risks due to their high mobility, as demonstrated by studies on aerosol deposition and coagulation behavior, as well as their influence on Hg levels in industrial urban areas. In contrast, thermally stable compounds like PbSO₄ and As₂O₅ exhibit lower bioavailability and reduced environmental reactivity. Experimental investigations across combustion temperature gradients enable the quantification of temperature-dependent migration rates, facilitating the optimization of combustion conditions—such as lowering temperatures to suppress the release of volatile metals. These findings provide a theoretical basis for targeted fly ash solidification technologies, supporting the precise immobilization of hazardous metal species.

A mechanistic understanding of the interplay between temperature and speciation enables actionable strategies for mitigating heavy metal emissions and enhancing pollution control frameworks in coal-fired power plants, as demonstrated by the adoption of clean coal technologies, improvements in energy efficiency, and the implementation of regulatory measures.

3.5. Ecological Risk Assessment of Heavy Metals Released from Raw Coal Combustion

The environmental hazards associated with heavy metal emissions from coal-fired boiler systems have emerged as a significant focus of scientific research over recent decades, as the ecological impacts of metal-contaminated slag and fly ash continue to pose persistent environmental challenges [55,56]. This understanding has driven the need for precise assessment methodologies to quantify the environmental risk potential of different heavy metal species, establishing such evaluations as a critical priority in the development of emission mitigation strategies for thermal power plants.

3.5.1. Enrichment Factor (EF) of Heavy Metals

The enrichment factor (EF) is a key indicator used to identify the sources of heavy metal elements in environmental matrices, as demonstrated by its widespread application in environmental geochemistry and pollution assessment. In this study, Al was selected as the reference element, and the EF values of heavy metals (Fe, Mn, Zn, Cu, Ti, Cd, Cr, Ni, Pb, As, Se, Hg) in slag (S1, S2) and fly ash (A1, A2) from two coal-fired power plants were calculated using Equation (4), as shown in Figure 4.

In slag S1, the EF values of Cu, Ti, and Hg were below 1, indicating minimal anthropogenic influence and suggesting that these elements are predominantly derived from natural sources. In slag S2, the EF values of Zn, Cd, Pb, As, Se, and Hg were also less than 1, implying a negligible contribution from coal combustion processes. Notably, Ni in slag S2 exhibited an EF value greater than 3, reflecting significant anthropogenic enrichment and posing potential environmental risks. For the remaining elements in both S1 and S2, EF values ranged between 1 and 3, indicating a moderate environmental impact attributable to coal-derived emissions.

According to the evaluation values presented in

Table 2. Thresholds for environmental factors (EF) and sources of pollution [31].

EF Range	Pollution Level	Source Interpretation
EF < 1	No enrichment	Natural sources
1 ≤ EF < 3	Mild enrichment	Natural or minor anthropogenic
3 ≤ EF < 10	Significant enrichment	Dominant anthropogenic (e.g., industry)
10 ≤ EF < 50	Strong enrichment	Clear anthropogenic origin
EF ≥ 50	Extreme enrichment	Severe pollution (e.g., mining)

. In fly ash A1, Cu (EF < 1) was primarily derived from natural sources. In fly ash A2, Cr and Hg (EF < 1) were predominantly of natural origin. Zn, Cd, and Pb in fly ash A2 exhibited EF values greater than 3, confirming their significant enrichment due to coal combustion. The remaining elements in both A1 and A2 showed EF values between 1 and 3, indicating minor but discernible environmental impacts attributable to coal combustion.

Krupnova et al.’s study on 23 metal elements in snow samples collected near coal-fired power plants revealed that the EF of As, Cd, Pb, and Zn exceeded 5 and exhibited a negative correlation with distance from the mining area, thereby confirming the significant contribution of coal combustion emissions to metal accumulation in the snow layer [57].

3.5.2. Geo-Accumulation Index (I_geo)

The geo-accumulation index (I_geo), a method developed by Muller in 1979, was employed to assess the contamination risks of pollutants by comparing the concentrations of heavy metals in the environment to their natural background levels. According to Equation (5), the calculated I_geo values for slag samples S1 and S2 were all below 1, indicating negligible contamination.

In fly ash A1, the I_geo value for Cd is 0.2, corresponding to low contamination risk. while that for Hg is 1.9, indicating a moderate contamination risk. In fly ash A2, the I_geo value for Cd is 1.6, classified as moderate contamination risk, and that for Hg is 0.7, classified as low contamination risk. For all other elements in fly ash A1 and A2, I_geo values were below 0, suggesting minimal to no contamination risk.

3.5.3. Long-Distance Dispersion of Heavy Metals from Raw Coal Combustion

The long-distance dispersion of heavy metals may pose significant risks to the surrounding environment and human health [58]. The Gaussian dispersion model (Equation (6)) was applied to estimate the concentrations of heavy metals at a horizontal distance of 1000 m from the two coal-fired power plants after one hour of continuous emission. Given that agricultural soils predominate beyond 1000 m, the calculated concentrations were compared with risk screening values for pH > 7.5, as specified in the “Soil Environment Quality standards for Agricultural Land Soil Pollution Risk Control (Trial)” (GB 15618-2018) [59]. Mn and Se were excluded from the assessment due to the absence of regulatory limits in the standard. Owing to their negligible volatility, Si, Fe, and Ti were also excluded; the analysis therefore focused on Zn, Cu, Cd, Cr, Ni, Pb, As, and Hg.

The theoretical modeling results presented in

Table 3. Long-range transport concentration of heavy metals.

HM	C1 (mg·kg−1)	C2 (mg·kg−1)	Risk Screening Value (pH > 7.5)	Risk Control Value
Zn	154.8	308.9	300	500
Cu	50.1	69.1	100	200
Cd	1.75	2.22	0.6	1.5
Cr	13.6	46.0	250	800
Ni	8.25	15.4	100	300
Pb	67.3	100.7	140	400
As	30.8	11.6	15	40
Hg	3.3	3.4	1	2

reveal distinct environmental risk profiles: emissions of Cu, Cr, Ni, and Pb from both C1 and C2 combustion processes remain below regulatory risk screening thresholds and control standards. In contrast, Cd and Hg concentrations significantly exceed both screening and control limits, identifying these elements as primary contaminants of concern. Notably, As emissions from C1 combustion substantially surpass screening thresholds, while Zn emissions from C2 combustion slightly exceed them. These modeled values correspond to individual combustion events. Current research indicates enhanced migration coefficients of Cd and Hg into fly ash phases for both coal types, along with elevated geoaccumulation indices (I_geo), suggesting increased potential for environmental exposure. The synergistic effects of their volatile speciation (e.g., CdCl₂, elemental Hg) and sustained concentrations in flue gas imply that chronic atmospheric deposition of Cd and Hg from coal combustion systems poses persistent and non-degradable ecological risks to adjacent agricultural ecosystems.

Studies have shown that As, Cd, Hg, and Pb exhibit similar distribution patterns in raw coal and in soils surrounding coal-fired power plants [60]. To further assess the potential long-range dispersion impacts of the two power plants on adjacent agricultural soils, soil samples collected beyond 1000 m were analyzed. The heavy metal concentrations were substantially below the risk screening thresholds, indicating effective retention of these elements by the flue gas treatment system. However, the efficient capture of volatile Cd and Hg species during coal combustion remains a significant challenge due to their high tendency to evade conventional retention mechanisms. In addition, we analyzed the concentrations of As and Cd in various tissues of wheat and vegetables (Chinese cabbage) cultivated in farmland surrounding the thermal power plant. The results indicated that the distribution of As and Cd in different wheat tissues followed the order: root > stem > grain. The As and Cd concentrations in wheat grains were below 0.1 mg·kg⁻¹. The concentrations in the edible parts (root and leaves) of Chinese cabbage were below 0.5 mg·kg⁻¹ for As and below 0.2 mg·kg⁻¹ for Cd. Both contaminants were within the maximum permissible levels specified in the “National Food Safety Standard—Limits of Contaminants in Food” (GB 2762-2022) [61].

To mitigate these risks, enhanced treatment of flue gas emissions—particularly targeting Cd, Hg, and As—is essential. Proven emission control strategies include the integration of activated carbon injection and fabric filters at emission outlets, as demonstrated by successful applications across various industrial sectors. Recent studies have confirmed the synergistic Hg removal capabilities of advanced air pollution control systems in coal-fired power plants [62], with technologies such as low-temperature electrostatic precipitators (LLT-ESP) and wet flue gas desulfurization (WFGD) systems exhibiting high efficiencies in Hg oxidation and removal. The implementation of these technologies would not only reduce atmospheric emissions but also alleviate long-term ecological pressures on surrounding agricultural ecosystems.

3.5.4. Potential Ecological Risk of Heavy Metals

The potential ecological risk index (RI) provides a systematic framework for quantifying composite toxic response coefficients of heavy metals resulting from coal combustion processes. This methodology enables a comprehensive assessment of three key environmental parameters: elemental enrichment dynamics, the leaching potential of specific contaminants, and integrated ecosystem risk profiles [63]. While conventional RI protocols incorporate eight pollutants—polychlorinated biphenyls (PCBs), Hg, Cd, As, Pb, Cu, Cr, and Zn—the present study focuses specifically on heavy metal-mediated ecological impacts and therefore excludes PCBs from the analytical framework.

The RI of heavy metals in slag and fly ash generated from the combustion of two raw coals, calculated using Equations (7) and (8), is summarized in

Table 4. Potential ecological risks of heavy metals in slag and fly ash.

		C1				C2
	${\mathit{T}}_{\mathit{r}}^{\mathit{i}}$	${\mathit{C}}_{\mathit{f}}^{\mathit{i}}$		${\mathit{E}}_{\mathit{r}}^{\mathit{i}}$		${\mathit{C}}_{\mathit{f}}^{\mathit{i}}$		${\mathit{E}}_{\mathit{r}}^{\mathit{i}}$
		S1	A1	S1	A1	S2	A2	S2	A2
Zn	1	0.20	0.43	0.20	0.43	0.03	1.25	0.03	1.25
Cu	5	0.16	0.40	0.80	2.00	0.64	0.83	3.20	4.15
Cd	30	0.69	1.70	20.70	51	0.32	4.60	9.60	138
Cr	2	0.11	0.22	0.22	0.44	0.42	0.16	0.84	0.32
Ni	5	0.13	0.27	0.65	1.35	0.42	0.29	2.10	1.45
Pb	5	0.25	0.52	1.25	2.60	0.09	1.10	0.45	5.50
As	10	0.09	0.35	0.90	3.50	0.01	0.16	0.10	1.60
Hg	40	0.14	5.60	5.60	224	0.57	2.36	22.8	94.4
RI				30.32	285.32			39.12	246.67

. As shown in the table, slag samples (S1 and S2) exhibited comprehensive RI values below 150, corresponding to a low-risk level, indicating minimal ecological risks associated with heavy metals in slag. Fly ash samples (A1 and A2) have RI values of 285.32 (A1) and 246.67 (A2), respectively, falling within the moderate-risk category, suggesting that heavy metals in fly ash may pose intermediate ecological hazards. These results highlight the differential ecological impacts of slag and fly ash, with fly ash exhibiting greater mobilization potential and bioavailability of heavy metals due to its finer particle size and higher reactive surface properties.

The potential ecological risk of heavy metals was determined in accordance with the ecological risk level standards presented in

Table 5. Ecological risk level standards [11].

${\mathit{E}}_{\mathit{r}}^{\mathit{i}}$	Pollution Degree	RI	Risk Degree
$E_r^i$ < 30	Slight	RI < 40	Slight
30 ≤ $E_r^i$ < 60	Medium	40 ≤ RI < 80	Medium
60 ≤ $E_r^i$ < 120	Strong	80 ≤ RI< 160	Strong
120 ≤ $E_r^i$ < 240	Very strong	160 ≤ RI< 320	Very strong
$E_r^i$ ≥ 240	Extremely strong	RI ≥ 320

. The ecological risk factors ($E_r^i$) for individual heavy metals in slag samples S1 and S2, as shown in Table 5 and Figure 5, were all below 40, indicating a low risk level. In fly ash sample A1, the $E_r^i$ value of Cd was 51, corresponding to a moderate risk level, while that of Hg reached 224, falling within the high risk category and signifying substantial ecological hazard. The ecological risk ranking of the eight heavy metals is as follows: Hg > Cd > As > Pb > Cu > Ni > Cr > Zn, with Hg posing the highest risk and Zn the lowest. In fly ash A2, the $E_r^i$ values of Cd and Hg were 138 and 94.4, respectively, both exceeding the threshold for elevated risk. The corresponding risk ranking is Cd > Hg > Pb > Cu > As > Ni > Zn > Cr. Overall, the elevated $E_r^i$ values of Cd and Hg in fly ash A1 and A2 are the primary contributors to their overall high ecological risks. To further evaluate the ecological impact of the two power plants on adjacent soils, soil samples were collected from areas at distances of 50–100 m, 100–500 m, and 500–1000 m. Analysis results revealed that Cd and Hg concentrations in the surrounding soils were significantly higher than those of other heavy metals. Moreover, Cd and Hg levels in the 50–100 m zone were markedly elevated compared to the 100–500 m zone, whereas concentrations in the 500–1000 m range were lower. Moskovchenko et al.’s study on the distribution of heavy metals in the snow layer surrounding coal-fired and oil-fired boilers revealed that the concentrations of elements such as Hg, As, Cr, Ni, and Zn were highest within a 0.5-km radius from the boiler sites [64].

Empirical investigations have confirmed that the RI for metallic elements in coal combustion facilities is 129.72, indicating a moderate ecological hazard classification. This quantitative assessment underscores the need for prioritized remediation strategies targeting Hg and Cd contamination, which are identified as the primary contributors to environmental degradation in these industrial systems [19]. Zhao et al. [65] systematically evaluated the environmental risks associated with coal combustion by-products (fly ash, slag, and gypsum), revealing that all three categories exhibit measurable environmental impacts. Specifically, fly ash poses moderate ecological risks for Hg, Cr, Cd, and Pb, while exhibiting elevated to extreme risks for As. Notably, Hg-containing by-products are susceptible to secondary release during storage, reuse, or leaching processes, thereby exacerbating potential environmental hazards [56]. Dai et al. [32] documented significantly elevated composite contamination indices in coal combustion byproducts (slag and fly ash), with Cd and Hg exhibiting ecological risk coefficients exceeding thresholds of 60 and 320, respectively, thereby posing substantial ecological risks to pedospheric environments. Charvalas et al.’s study revealed that the pollution levels of ten potentially toxic elements (PTEs) in three urban areas adjacent to the industrial zone of Volos were at moderate levels, indicating that industrial activities may exert a carry-over effect on soils in residential areas [66]. Of particular concern is the high contamination potential of these heavy metals, which has significant implications for soil matrix stability and biota viability. The method proposed by Hakanson was adopted in this study due to its widespread application in heavy metal pollution assessments. However, in recent years, scholars have introduced new approaches aimed at refining and improving Hakanson’s methodology. Leshukov et al. conducted a study on dust deposited in the snow layer during the snow season at varying distances from coal mines and coal-fired power plants in the Kemerovo region (Russia). The findings indicate that particulate matter generated during coal mining and the operation of coal-fired power plants may pose adverse health risks to the local population in this region [67]. Ramires et al.’s study indicates that Mn, Fe, Pb, As, Cr, Ni, Zn, and Cu are the primary elements contributing to human health risks [68]. Dinesh et al. state that Mn, Cu, Zn, Pb, Cr, Ni, As, Cd, Se, and Hg are the primary elements emitted from coal combustion and industrial activities [69]. Zhuravleva et al.’s study indicates that As, Be, Cl, Co, Cr, F, Hg, Mn, Ni, V, Pb, Se, Ti, and Zn are the elements of primary concern for monitoring in coal dust [70]. These studies have improved the evaluation of heavy metal pollution risks through methodological advancements.

To mitigate emissions of Cd, Hg, and other heavy metal contaminants during coal combustion, an integrated approach combining pre- and post-combustion control strategies should be prioritized. Pre-combustion interventions involve coal beneficiation technologies, including advanced coal washing processes, which effectively remove inherent metal-bearing mineral components from raw coal. Post-combustion mitigation employs adsorption—based flue gas purification systems, in which the injection of high-surface-area adsorbents—such as limestone (CaCO₃), calcium oxide (CaO), and bauxite-derived materials—has proven effective in capturing volatile heavy metal species from combustion flue gases [71]. Heavy metals can be systematically removed using existing pollution control systems in coal-fired power plants, such as electrostatic fabric composite dust collectors and flue gas desulfurization units [5]. Furthermore, new pollution control technologies can be developed, and combustion conditions can be optimized through measures such as temperature regulation and enhancement of desulfurization reactions.

Emissions of Cd and Hg during coal combustion pose significant environmental and health risks due to their extensive spatial dispersion, complex speciation dynamics, and potential for long-term bioaccumulation. The chemical speciation of heavy metals in the environment serves as a key indicator for assessing their toxicity. Further investigation into the speciation of Hg and Cd—elements identified in this study as posing significant ecological risks—will be a primary focus of future research. Targeted remediation measures, including engineered controls such as soil stabilization and ecological restoration techniques like phytoremediation using metal-accumulating plants (e.g., Brassica), are essential for mitigating soil metal enrichment. In this study, Power Plant 1 employs an external flue gas emission, necessitating scheduled monitoring of exhaust emissions and timely upgrades to flue gas treatment equipment. In contrast, Power Plant 2 utilizes an internal emission containment system, which requires continuous monitoring of indoor pollutant concentrations to reduce occupational health risks for workers. If Power Plant 2 were to transition to external flue gas emissions, it would be required to implement the same monitoring protocols and equipment enhancement measures currently in place at Power Plant 1.

This research outcome is based on the characteristics of Shanxi Province, which has abundant coal reserves and high coal consumption. The release behavior of 13 heavy metal elements from raw coal to slag and fly ash during the combustion process, as well as the assessment of potential ecological risks in this study, can be extended to regions with substantial coal production and similar coal quality for predicting heavy metal emissions during coal combustion.

4. Conclusions

The study revealed notable differences in heavy metal migration behavior between the two coal-fired plants. Facility C1 exhibited higher release rates of Cr, Ni, Se, and Hg compared to C2, whereas C2 showed greater migration of Mn, Cu, Ti, Cd, Pb, and As. Analysis indicated that the elevated metal content in raw coal at C1 was proportionally enriched in both slag (S1) and fly ash (A1), while at C2, metals were preferentially partitioned into slag (S2) with minimal accumulation in fly ash (A2). Cd, As, Pb, and Hg demonstrated a higher tendency to migrate to fly ash in both facilities relative to other heavy metals.

The release of heavy metals (Cd, As, Pb, and Hg) during coal combustion is influenced by their chemical speciation. Although raw coals C1 and C2 exhibit different phase distributions, their shared coalification characteristics result in similar transformation patterns. (1) Cd is primarily bound to organic matter in coal, converting into residual fractions in slag but persisting in multiple phases in fly ash. (2) As mainly exists in sulfide-bound and residual forms in coal and is predominantly transformed into residual fractions after migration. (3) Pb is distributed across all four fractions in coal, showing minimal retention as a residual fraction in slag but predominating in fly ash. (4) Hg is mainly associated with sulfides and organic matter in coal, transforming into Fe/Mn oxide-bound and residual fractions in slag, and into sulfate-bound and residual forms in fly ash.

Combustion temperature significantly influenced heavy metal concentrations in flue gas emissions. Under progressively increasing thermal conditions, the release rates of Cd, As, Pb, and Hg from both coal types exhibited nonlinear increasing trends. Hg, due to its high volatility, showed particularly pronounced migration behavior, with approximately 90% of its content released from the raw coal matrix at 600 °C. Comparative analysis revealed no significant difference in the migration behavior of these four elements between the two coal types.

The release of Cd, Hg, and As during coal combustion in power generation facilities poses significant potential ecological risks, with Cd and Hg exhibiting the most severe contamination effects. Proactive mitigation strategies—such as the implementation of advanced flue gas purification systems and the establishment of real-time emission monitoring systems—are essential measures for risk reduction and the protection of the surrounding ecosystems.