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Environ Eng Res > Volume 30(2); 2025 > Article
Lee, Kim, Choi, Park, Lee, Shin, Han, Back, Hong, Yoon, and Yun: Mercury emission and fate characteristics in various combustion sources

Abstract

In this study, we investigated the behavior and emission characteristics of Hg across diverse industrial combustion sources, such as coal-fired power plants, solid refuse fuel (SRF) power plants, medical waste incinerators, and industrial waste incinerators, as well as the development of emission factors, Among the facilities, the estimated Hg control efficiency was 86% for coal-fired power plants, 69% for SRF power plants, and over 95% for medical and industrial waste incinerators. The oxidation and regulation of elemental Hg (Hg0) are considered important factors in reducing Hg air emissions, with the flue gas HCl concentration being the primary factor affecting Hg oxidation. Particulate Hg was mainly controlled in the electrostatic precipitator (ESP) at the power plant, effectively capturing dust particles ranging from 10 to 100 μm. The emission factors estimated by measuring stack flue gas exhibited their highest values at SRF power plants, with an estimated average of 127±25 mg/ton, while the lowest values were observed for industrial waste incinerators, with an average of 2.5±0.3 mg/ton. This study is significant in that it provides a comparative analysis of various real industrial cases, and the Hg emission factors are expected to offer valuable data for estimating future national Hg air emissions.

Graphical Abstract

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Introduction

Hg is a toxic heavy metal present ubiquitously in the global ecosystem. It is globally circulated through various media, adversely affecting human health and the environment. The United Nations Environment Programme (UNEP) convened the Minamata Convention on Mercury in 2013 with the primary goal of mitigating Hg emissions and safeguarding human health and the environmental integrity [1]. The convention came into effect on August 16, 2017. Article 8 pertains to the control and reduction of Hg emissions. To establish a feasible strategy, the primary emission sources are listed in Annex D, including coal-fired power plants, coal-fired industrial boilers, nonferrous metal production plants, waste incinerators, and cement clinker production facilities. In addition, Annex 11 addresses the classification and management of Hg waste, categorizing it as Hg-containing waste, Hg-consisting waste, and Hg-contaminated waste.
Global inventories for Hg emissions into the air from human sources have been compiled at approximately five-year intervals since 1990 by scientific expert groups [2]. According to an inventory conducted in 2018, approximately 2220 tons of Hg are emitted into the atmosphere annually. Among the various emission sources, artisanal and small-scale gold mining constitutes the largest proportion with 838 tons of Hg (37.7%). Coal-fired power plants come in second, emitting 292 tons of Hg (13.1%), while Hg air emissions from waste incineration have steadily increased, reaching a total of 162 tons of Hg (7.27%) annually.
Hg emitted into the air can be classified into three categories: particulate Hg (Hgp), oxidized Hg (Hg2+), and elemental Hg (Hg0). These forms of Hg, present in combustion off-gases, exhibits distinct behaviors within an air pollution control device plant, depending on their respective physicochemical and mechanical characteristics. Therefore, previous studies investigated the control characteristics of Hg for various air pollution control devices (APCDs), such as selective catalytic reduction (SCR), electrostatic precipitator (ESP), and flue gas desulfurization (FGD). The initial coal-fired power plants consisted solely of ESP and FGD as APCDs, resulting in a mere 50% control efficiency for Hg emissions [3]. Therefore, the stack emissions of Hg were relatively high. This can be attributed to the fact that while some Hgp and Hg2+ are effectively controlled, Hg0 is emitted into the air. The localized increase in broiler temperature to approximately 1500°C was achieved by utilizing coal burners [4]. Moreover, the Hg contained in the fuel is converted into a Hg0 at temperatures exceeding 800°C [5]. Subsequently, the effect of applying SCR to reduce nitrogen oxides (NOx) on the control mechanisms for Hg was studied, and it was reported that SCR has a significant Hg0 oxidation effect of more than 80%, with the majority of oxidized Hg subsequently being absorbed and removed through a scrubber in the FGD system [6]. Furthermore, the recent inclusion of APCDs, such as SCR, CS-ESP, and FGD, in coal-fired power plants has resulted in additional benefits, including enhanced Hg removal and improved optimization of facilities for controlling NOx, sulfur oxides (SOx), and particulate matter (PM) [7].
Recently, the types of waste have diversified owing to lifestyle changes and rapid industrial growth. In addition, as the amount of incinerated waste continues to rise, Hg air emissions inevitably follow suit. However, waste incineration facilities employ a standardized sequence of APCDs in the following order: semi-dry reactor (SDR), fabric filter (FF) with activated carbon injection (ACI), and scrubber. However, it is worth noting that the application of SCR faces limitations as it significantly affects the oxidation and removal of Hg. Most waste incineration facilities are stoker type and are operated at temperatures in the range of 800–1000°C [8]. In such facilities, the Hg contained in waste is converted into gaseous or particulate Hg, and Hg0 is at risk of being emitted into the atmosphere without undergoing treatment via the SCR process. Additionally, factors such as, NO, SO2, and halogen substances in the flue gas, play significant roles in affecting Hg oxidation. Hg0 has an oxidation efficiency of approximately 40% under conditions of 600 ppm NO and 13 ppm chlorine gas (Cl2) [9]. Therefore, the effectiveness of Hg oxidation is significantly influenced by the technical characteristics of APCDs and the composition of waste. This was further corroborated by a study that examined Hg species emitted from municipal solid waste incinerators in China [10]. The analysis revealed a significant variation, ranging from 7.2 to 93.2 μg/Nm3, with Hg2+ constituting approximately 90% of the common ratio. The high chloride content in the waste is considered the primary factor contributing to this phenomenon. A study investigating Hg species and mass balance for solid refuse fuels (SRF) power plants estimated the Hg content in boiler off-gas to be 32.6 μg/Sm3, ranging from 30.5 to 216.7 μg/Sm3 [11]. Unlike municipal solid waste, SRF is considered a fuel, and both heavy metals and chlorides are carefully managed during the manufacturing process. Despite the advantage of having a low Hg concentration, Hg0 constitutes approximately 50% of the emissions because of the relatively low chloride content in fuels.
With the increase in facilities and the diversification of combustion sources, the behavior of Hg has become increasingly complex, resulting in rising emissions. For example, in addition to coal-fired power plants and industrial waste incinerators that emit significant amounts of mercury, there has been a recent increase in SRF power plants and medical waste incinerators. Therefore, gaining a deep understanding of Hg behavior across various combustion sources is of paramount importance for reducing Hg air emissions and enhancing control efficiency. In this study, we aimed to identify the behavioral characteristics of Hg in coal-fired power plants, medical waste incinerators, SRF power plants, and industrial waste incinerators—representative sources of atmospheric Hg emissions. To the best of our knowledge, this study is the only field case that quantitatively compares mercury emissions across the four different industries mentioned above. Additionally, Hg emission factors were developed for evaluating each facility, providing valuable academic data for estimating future national Hg air emissions.

Materials and Methods

2.1. Summary of Various Combustion Sources and APCDs

For analyzing Hg behavior and mass balance, several industrial facilities, such as coal-fired and SRF power plants and industrial and medical waste incinerators, were selected. Isokinetic sampling was conducted both upstream and downstream of each APCD as well as in the stack of each facility [12]. A summary of these four facilities is presented in Table 1.
To study Hg behavior, we selected a coal-fired power plant that uses pulverized bituminous coal (blending coal) as its power source and has a power generation capacity of 500 MW. This power plant employs typical APCDs, including SCR for NOx reduction in combustion gases, utilizing NH3 as a reducing agent. Additionally, a cold-side ESP (ESPc) is used to control particulate matter, and a subsequent FGD process induces a desulfurization reaction using lime/limestone.
The selected SRF power plant has a power generation capacity of 10 MWh and a daily consumption of approximately 80 tons/day of refuse-derived and plastic fuel. This facility consists of a selective non-catalytic reduction (SNCR) unit, an SDR, and an FF with an ACI for heavy metal control.
The selected stoker-type industrial waste incineration facility has an incineration capacity of 48 tons/day. It employs an APCD configuration similar to that of the SNCR-SDR-FF with an ACI. This facility incinerates waste at a ratio of one-half industrial waste to municipal solid waste.
Lastly, the stoker-type medical waste incineration facility has an incineration capacity of approximately 40 tons/day, and it utilizes SNCR-SDR-FF with an ACI-scrubber as an APCD. In this facility, medical waste mixed with COVID-19 waste is injected into a boiler.

2.2. Sampling and Hg Analysis

Samples were collected from each process, including the inflow materials and fuels, to estimate the overall flow and behavior of Hg. Mercury sampling was performed in triplicate to enhance the reliability of the data at both upstream and downstream sections of the APCD system. The Ontario–hydro method was used to analyze Hg chemical species, such as Hgp, Hg2+, and Hg0, in the flue gas [13], which is a highly reliable method adopted by the US EPA. Isokinetic sampling was performed to collect Hgp, and samples ranging from 1.0 to 2.5 Sm3 were collected for 2–3 h. The collected gaseous samples were passed through eight impingers, with gaseous Hg2+ and Hg0 being collected in each solvent. Impingers 1–3 captured Hg2+ and contained a 1 N KCl solution. Impingers 4–7 collected Hg0 using 5% HNO3/10% H2O2 and 10% H2SO4/4% KMnO4. The final impingement contained silica gel. The collected samples were subjected to a pretreatment process and were analyzed using an Hg analyzer (RA915+ ZEEMAN, Lumex Ltd, St. Petersburg, Russia) employing the cold vapor atomic absorption spectroscopy (CVAAS) method. To ensure the accuracy of Hg concentration analysis, the determination coefficient (R2) of the calibration curve was maintained at 0.99 or higher. Among the Hg species, the particle-phase Hgp was collected using an 88R thimble filter (Advantec, Tokyo, Japan). The flue gas composition, including O2, CO, CO2, NO, NO2 and Sox, was analyzed at each sampling point using an MK-9000 analyzer (ECOM GmbH, Nordrhein-Westfalen, Germany). The feedstock fuel and waste were analyzed to ensure fuel quality and inspect for any potential pollutant generation. Elemental analysis was conducted to examine the C, H, O, N, and S content using an EA1110 elemental analyzer (Thermo Finnigan Co., San Jose, CA, USA). Proximate analysis was conducted using a thermogravimetric analyzer (TGA701, LECO Co., St. Joseph,. MI, USA). In addition, the chlorine content in the sample was determined using ion chromatography (ICS-2100, Dionex Co., Sunnyvale, CA, USA).
The solid and liquid samples were pre-treated using the US EPA Methods 7471a and 7470a, respectively [14] [15]. For the pre-treatment, 0.2 g of solid samples and 100 mL of liquid samples were placed in separate borosilicate brown bottles. Aqua regia and sulfuric acid were added to dissolve the samples. Hg oxidation and reduction were carried out using 15 mL of a 5% (w/v) KMnO4 solution and 6 mL of a 12% (w/v) SnCl2 solution.

2.3. Estimation of Hg Emission Factor

The annual amount of air pollutants was calculated using the emission factor developed based on field measurement data. Air pollutant emissions were estimated by multiplying the emission factors by fuel consumption. The emission factor was calculated using the following equation (Eq. (1)):
(1)
Emission factor=Pollutantconcentration(g/m2)×Gasflow(Sm3/day)Fuelconsumption(ton/day)

Results and Discussion

3.1. Fundamental characteristics of fuels and waste

A comparison of the proximate analyses of bituminous coal and SRF indicated that the volatile components of the SRF were higher (Table S1 in the supplementary materials). Volatile matter directly affects the heating value and fuel quality because it is associated with the generation of air pollutants during fuel combustion. Aich et al. reported the fixed carbon content of fuels affects the initial combustion performance and later stabilization [16]. They found that the combustion performance increased within the range of 22wt% to 34wt% in the fixed carbon content. This blended coal has a slightly higher fixed carbon content of 50wt%, which can be adjusted by mixing it with bio-SRF to achieve the desired combustion conditions. In contrast, the fixed carbon content of the SRF is typically less than 5%, which could potentially lead to the formation of thermal NOx and a significant increase in the initial boiler temperature. The elemental analysis revealed that C, H, N, and O were the major components, corresponding to volatile matter and moisture content. The S content values in coal and bio-SRF were 0.4wt% and 0.9wt%, respectively, while the volatile content of SRF was less than 0.05wt% of the detection limit. This highlights the need for more intensive control of NOx concentrations in the flue gas of SRF power plants.
Heavy metals, such as Cr, Cu, and Ni, which have relatively high boiling points, were found to remain in the bottom ash [17]. In addition, Cd and Pb, which volatilize at high temperatures, were distributed in the form of fly ash and effectively controlled using fabric filters. Evans and Williams reported that municipal solid waste typically contains heavy metal concentrations ranging from 10 to 40 ppm of Cd, 100 to 450 ppm of Cr, 450 to 2500 ppm of Cu, 50 to 200 ppm of Ni, and 750 to 2500 ppm of Pb [18]. However, SRF exists in the form of pellets and exhibits a relatively low heavy metal concentration owing to inorganic component screening and high-temperature treatment at approximately 600°C, during the manufacturing process. Therefore, the quality of both blended coal and SRF used in this study was excellent.
The proximate analysis data for both types of waste are presented in Table S2 in the supplementary materials. Their ash fractions were similar, but the moisture content was twice as high in the medical waste, reaching approximately 28%. This higher moisture content is due to medical waste containing fluids and tissues. Additionally, the carbon content is higher in industrial waste, which may lead to relatively low initial incineration efficiency for medical waste. Furthermore, since many components of medical products are synthetic resins, the chlorine (Cl) content was significantly high at 1.89%. Cl indirectly affects the oxidation of Hg and can enhance the efficiency of Hg control in the flue gas.

3.2. Hg Speciation from Various Combustion Sources

3.2.1. Coal-fired power plant

Most of the Hg in the fuel undergoes thermal decomposition, transitioning into the gaseous phase when exposed to the high-temperature zone of the boiler. Control measures for gaseous Hg are limited to adsorption and absorption. However, controlling Hg0 depends on physical adsorption and is known to be relatively inefficient. To effectively reduce air emissions of Hg, improving the conversion of Hg2+ is essential, and the co-beneficial effect of arranging APCDs becomes crucial.
The Hg concentrations in the APCD stream flue gas of the CPPs are showed in Fig. 1. The Hg concentrations upstream and downstream of the CPPs were 5.23 and 0.7 μg-Hg/m3, respectively, with an overall control efficiency of 86.3%. The study of Senior et al. indicated that the equilibrium of Hg species in the flue gas depends on temperature conditions [19]. At temperatures of approximately 800°C or higher, Hg was mostly converted to Hg0. The Hg concentration at the gas inlet of the SCR was 3.79 μg-Hg/m3, which reduced to 1.86 μg-Hg/m3 after passing through the SCR process, with the catalyst estimated to have an oxidation efficiency of approximately 51%. Furthermore, after passing through an ESP, the Hg0 concentration was reduced to at least 0.4 μg-Hg/m3. Finally, the FGD process showed that the remaining Hg2+ and Hgp in the flue gas were mostly controlled. Some FGDs systems have reported re-emission of Hg due to limestone use and pH changes, but these effects were found to be negligible. The predominant factors affect Hg oxidation are catalysts or electrostatic forces, and the Hg0 concentration dramatically decreases after passing through these plants. In addition, gaseous components, such as HCl, NOx, and SOx, play a complex role in Hg oxidation.

3.2.2. SRF power plant

The Hg concentrations in the APCD stream flue gas of the SRF power plant are listed in Fig. 2. Field tests showed that the Hg0 concentration of the inlet and outlet APCD gases reached 15.6 and 11.0 μg-Hg/m3, respectively, with an oxidation efficiency of only 29.5%. In addition, the Hg0 concentration in flue gas was found to be as high as 203 μg-Hg/m3 due to SRF heterogeneity and raw waste material characteristics. Despite the pre-treatment process in SRF manufacturing, there is significant deviation in Hg concentration that requires careful management. Compared with CPPs, the concentration of Hgp collected by isokinetic sampling in the SRF power plant was significantly higher at 14.4 μg-Hg/m3. Li et al. reported the concentration of Hg0 increased dramatically at HCl concentrations ranging from 0 to 10 ppm [20]. However, it gradually decreased as the HCl concentration surpassed 20 ppm. The on-site testing results showed that the HCl concentration in the upstream gas was 12 ppm, considered suitable for Hg oxidation. SO2 proved effective for Hg oxidation when O2 levels were maintained at 4% or when the HCl concentration exceeded 5 ppm in the flue gas. Although the oxidation efficiency of Hg at approximately 100 ppm of SO2 was estimated to be 60%, in the SDR, Hg2+ was re-emitted because of the chemical reaction. Because SRF power plants do not include NOx or Hg0 oxidation in their APCDs, the Hg oxidation efficiency varies depending on the gas composition effects.

3.2.3. Medical waste incinerator

The Hg concentrations in the ACPD stream flue gas of the medical waste incinerator are listed in Fig. 3. The Hg concentrations of the upstream and downstream APCDs of the medical waste incinerator were 52.5 and 0.8 μg-Hg/m3, respectively, with almost all of the Hg being effectively removed by APCDs. The concentration of NOx in the boiler flue gas averaged 16.5 (7.6 to 59) ppm, and the HCl concentration was significantly higher at 724 ppm. The elevated HCl levels originated from Cl-containing materials in medical waste, polymers, and plastics used for storage and contribute to the oxidation of Hg. Consequently, despite maintaining the incinerator temperature above approximately 950°C, the fraction of Hg0 remained low. In addition, Hg2+ was reduced to 2.3 ppm in the SDR, which was controlled by absorption at approximately 77%. The SDR acted at the first APCD, a somewhat unusual step, significantly controlling Hg in the boiler downstream gas. The dominant factor for Hg oxidation in flue gas was the high HCl concentration, which remained at 102 ppm downstream of the SDR, contributing to the oxidation of persistent Hg0. Most of the total PM and Hgp in the flue gas were removed by the bag filter. As shown in Fig. 3, the bag filter efficiently removed most of the TPM and Hgp in the flue gas, while HCl was reduced to approximately 10 ppm through absorption in the wet scrubber solution.

3.2.4. Industrial waste incinerator

The Hg concentrations in the APCD steam flue gas from the municipal solid waste incinerator are shown in Fig. 4. The Hg concentration of the APCDs upstream and downstream of the municipal solid waste incinerator were 24.3 and 0.2 μg-Hg/m3, respectively. The concentration of NOx in the boiler flue gas averaged 19 (ranging from 5 to 59) ppm, while the HCl concentration was lower than that of the medical waste incinerator, measuring at 2.9 ppm. The NO concentration in the industrial waste incinerator was similar to that in the medical waste incinerator, but the HCl concentration was significantly lower. The type and composition of the input waste influenced changes in Hg speciation characteristics downstream of the boiler. Half of the Hg generated from industrial waste combustion is in the form of Hgp, with an Hg0 concentration of 8 μg-Hg/m3, which is higher than the Hg2+ concentration of 4.9 μg-Hg/m3. When the Hg in the flue gas was subjected to SDR, 77% of the Hgp and 71% of the Hg2+ were effectively controlled. Traditional waste incinerators are typically configured with APCDs, with SDR considered as the most efficient unit for Hg control. The control efficiency of Hg was estimated to be approximately 70% and was further enhanced as the proportion of Hg2+ increased, primarily because of the effect of acidic gases, including HCl, downstream of the boiler. However, Hgp and Hg0 concentrations remained relatively high at 3.0 and 8.7 μg-Hg/m3, respectively. The Hg Downstream of the bag filter, Hgp was effectively removed, and Hg0 was controlled at approximately 75% efficiency through heterogeneous physicochemical adsorption within the pores of fly ash [21]. Finally, most air pollutants, such as HCl and persistent Hg, were effectively controlled using wet scrubbers.

3.3. Characteristics of Hgp and Residues

The characteristics of fly ash affect the adsorption of Hg in the flue gas and its concentration in air emissions. Fly ash has various factors that affect the physicochemical adsorption of Hg, including unburned carbon, surface area, surface morphology, and oxidation potential. This section discusses the correlation between particle size and Hg in fly ash from various combustion sources. Fig. 5 show the size distribution of fly ash in thermal power plants and waste incineration facilities.
The particle size distributions of the combustion residues have a direct correlation with the heavy metal content via adsorption. The average particle sizes of coal fly ash (F/Acoal) and SRF fly ash (F/ASRF) were estimated to be 109.57 and 57.95 μm, respectively. F/Acoal has an average diameter and standard deviation approximately twice those of F/ASRF. Koukouzas et al. reported that fly ash with a small particle size has a large surface area [22]. F/Acoal accounted for approximately 6.5% of particles smaller than 1 μm, as determined by ESPc. In addition, the Brunauer-Emmett-Teller (BET) surface area was estimated to be 3.44 m2/g. Wang et al. reported a surface area ranging from 4.0 to 21.2 m2/g and pore volume ranging from 4.89×10−3 to 2.01×10−2 for fly ash from various industries. These fly ash types have relatively small surface areas, with F/ASRF having a higher surface area than F/Acoal because it is mixed with bag filter dust and injected with activated carbon in silo. The adsorption–desorption isotherms are shown in Fig. 6, indicating that F/ASRF had a greater adsorption capacity than F/Acoal. Furthermore, the heavy metal concentrations in F/ASRF were found to higher than those in F/Acoal. Ahmad et al. reported the findings of a study on the distribution of heavy metals and substances in municipal solid waste incineration facilities [17], which indicated that, similar to previous studies, Cd, Pb, and Hg were the major substances in F/ASRF, while Cr, Cu and Ni were the primary heavy metal components in B/ASRF.
The average particle sizes of medical waste fly ash (F/AMW) and industrial fly ash (F/AIW) were determined to be 9.65 and 5.51 μm, respectively. The diameter of the fly ash particles in the waste incinerator was significantly smaller than that in the coal-fired power plant. Coal is a relatively homogeneously sized solid fuel that is roughly pulverized during storage and feeding. In the SCR process, larger ash particles are physically filtered out, and the electrostatic effect controls fly ash particles in the size range of approximately 10 to 100 μm. In contrast, waste inputs to waste incinerators are highly heterogeneous in size and composition, resulting in the generation of a large range of particle diameters during combustion. The bag filter collects these particles through inertial impact and diffusion, and it is possible to control ultrafine dust particles if a dust layer forms on the fabric surface [23]. In addition, powdered activated carbon for heavy metal control was injected near the bag filter, resulting in a relatively high amount of ultrafine dust. Therefore, the concentration of Hgp generated from waste incineration facilities is slightly high, and most of it is discharged in the form of waste, such as fly ash, necessitating proper management.

3.4. Comparison of Hg Emission Factors

In this study, the air emission factor of Hg was estimated, and measurements of Hg were conducted downstream of various facilities (Eq. (1)). The calculated emission factors for these facilities are as follows: coal fired-power plant 14.98±2.4 mg/ton, SRF power plants 127±25 mg/ton, medical waste incinerator 10.8±2.2 mg/ton, and industrial waste incinerator 2.5±0.3 mg/ton (Fig. 7). These Hg emission factor values were used to estimate the total national Hg emissions from the respective source categories. Previous studies have reported emission factors ranging from 0.1 to 0.3 g/ton [11] and 17.62 mg/ton [9]. The emission factor was reduced by approximately one-twentieth compared to that at the beginning of the study. The introduction of SCR-ESPc-FGD as the best available technology for coal-fired power plants is expected to significantly reduce Hg emissions in commercial facilities. Additionally, regulations regarding production and emission control have been strengthened. Similarly, the Hg emission factor for waste incinerators was reduced by approximately a quarter, from 47.2 to 10.8 mg/ton. On the other hand, as shown in Fig. 7, the SRF power plant exhibited an exceptionally high Hg emission factor, which is presumed to be due to the absence of a wet scrubber. The Hg emission factors developed in this study can be utilized as a current data source for estimating national Hg air emissions from various combustion sources through the Mercury Inventory Toolkit provided by the UNEP.

Conclusions

By measuring Hg emissions from various combustion sources, this study achieved two objectives. Firstly, it identified the behavior and control characteristics of particulate and gaseous Hg in the APCDs of various industrial facilities. The oxidation of Hg0 was found to have the most significant impact on reducing Hg emissions into the atmosphere. Coal-fired power plants exhibited an Hg control efficiency of 86%, SRF power plants approximately 69%, and waste incineration facilities approximately over 95%. Furthermore, concerning particulate Hg, the combination of inertial impact and diffusion within the bag filter, along with the formation of a dust layer, allowed for some degree of control, even for ultrafine dust. In addition, among the acidic gases present in the flue gas and within the prevention facility, HCl emerged as the most dominant influencing factor. The concentration of HCl at the end of the boiler of the medical waste incinerator was significantly high at 724 ppm, contributing significantly to the oxidation of Hg0. Secondly, emission factors for Hg were developed for each industrial facility. The SRF power plant had the highest emission factor of 127±25 mg/ton, while the industrial waste incinerator had the lowest emission factor of 2.5±0.3 mg/ton. These Hg emission factors were estimated based on the Hg content in the fuel and the control efficiency of the air pollution prevention facility. The limitation of this study is that removal efficiencies could not be equally compared due to differences in facility scale and process efficiency; therefore, future research will include quantitative studies to address and compensate for these differences. Nevertheless, the academically derived Hg emission factors will serve as a foundation for future estimates of national Hg air emissions.

Supplementary Information

Acknowledgement

This research was supported by the Korea Environmental Industry & Technology Institute (KEITI) and the Ministry of Environment of the Republic of Korea (Grant No. 2021003350010).

Notes

Conflict-of-Interest Statement

The authors declare that they have no conflict of interest.

Author Contributions

E.L. (Post-Doctoral Researcher) conducted the experiments and wrote this manuscript (First author). J.K. (Senior Engineer) analyzed the data. Y.C.(Researcher) and K.P.(Researcher) conducted the experiments and analysis. C.L. (Principal Engineer) analyzed the data. W.S. (Senior Researcher), S.H. (Principal Engineer), and S.B.(Researcher) revised the manuscript. W.H.(CEO) organized the project. H.Y. (Senior Researcher) and J.Y. (Principal Researcher) conducted writing-review & editing; and supervision (co-corresponding authors).

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Fig. 1
Hg speciation across air pollution control devices (APCDs) of the coal-fired power plant.
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Fig. 2
Hg speciation across air pollution control devices (APCDs) of solid refuse fuel (SRF) power plant.
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Fig. 3
Hg speciation across air pollution control devices (APCDs) of medical waste incinerator.
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Fig. 4
Hg speciation across air pollution control devices (APCDs) of the industrial waste incinerator.
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Fig. 5
Size distribution of fly ash in thermal power plant and waste incineration facilities. (a) Particle size distribution of fly ash from industrial waste incinerator, (b) Particle size distribution of fly ash from medical waste incinerator, (c) Particle size distribution of fly ash from coal-fired power plant, (d) Particle size distribution of fly ash from SRF power plant.
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Fig. 6
Adsorption and desorption isotherms of fly ash from (a) SRF power plant and (b) coal.
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Fig. 7
Calculated Hg emission factor from various combustion sources (n=3).
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Table 1
Summary of various combustion sources and air pollution control devices (APCDs)c
Combustion source Type of APCD Description
Facility Type of boiler Capacity
Coal-fired power plant fluidized bed 500 MW SCR + ESPc + FGD Bituminous coal (with Bio SRF)
SRF power plant Circulating fluidized bed 10 MW SDR + FF(ACI) RDF + RPF (Pellet type)
Industrial waste incinerator Stoker 48 tons/day SDR + FF(ACI) + Scrubber Industrial:municipal waste = 5:5
Medical waste incinerator Stoker 40 tons/day SDR + FF(ACI) + Scrubber Medical waste with COVID-19 waste
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