^a MOE Key Laboratory of Thermo-Fluid Science and Engineering, Xi’an Jiaotong University, Xi’an 710049, China; ^b School of Ecology and Environment, Hainan University, Haikou 570228, China; ^c Faculty of Mechanical Engineering &Naval Architecture, University of Zagreb, Zagreb 10000, Croatia; ^d Faculty of Mechanical Engineering, University of Maribor, Maribor 2000, Slovenia; ^e Faculty of Mechanical Engineering, University of Belgrade, Belgrade 11120, Serbia; ^f AGH University of Krakow, Krakow 30059, Poland; ^g Consortium for Clean Coal Utilization, Department of Energy, Environmental & Chemical Engineering, Washington University in St. Louis, St. Louis, MO 63130, USA

In response to the current concern regarding global warming, countries across the world signed the Paris Agreement in 2016, with the consensus of limiting the global average temperature rise to 2.0 or {1.5}^{\circ } \mathrm{C} compared with pre-industrial levels [1]. The widespread use of fossil fuels has resulted in massive emissions of {\mathrm{C}\mathrm{O}}_2, which is considered to be one of the primary causes of global warming. While it is necessary to reduce carbon emissions, it is difficult to stop using fossil fuels in the short term, as they are low-cost, reliable sources of energy. As a result, there is a need to further develop carbon capture, storage, and utilization (CCUS), which employs three main technologies: ① Pre-combustion capture, which typically involves an integrated gasification combined cycle system in which the carbon in the fuel is gasified as \mathrm{C} \mathrm{O} and then transformed into {\mathrm{C}\mathrm{O}}_2, so as to be captured; ② post-combustion capture, in which coal combustion occurs in a traditional air atmosphere, and the {\mathrm{C}\mathrm{O}}_2 in flue gas is separated via physical and chemical methods; and ③ oxy-combustion, wherein coal is combusted with a mixture of pure {\mathrm{O}}_2 and recirculation flue gas, producing high-concentration {\mathrm{C}\mathrm{O}}_2 at the outlet that can be directly captured [2]. Among these, oxy-combustion is regarded as the most promising technology for large-scale power plants.

Oxy-combustion was first proposed for the production of {\mathrm{C}\mathrm{O}}_2 for enhanced oil recovery (EOR) [3]. A simplified diagram of oxy-combustion is shown in Fig. 1 [3]. The adiabatic flame temperature for pulverized coal (PC) burning in pure {\mathrm{O}}_2 can be significantly higher than that of traditional air combustion. Therefore, some of the flue gas is recycled back into the furnace to reduce the flame temperature [4] in a process also referred to as {\mathrm{O}}_2 / {\mathrm{C}\mathrm{O}}_2 combustion. The fundamental difference between oxy-combustion and air combustion is that {\mathrm{C}\mathrm{O}}_2 replaces {\mathrm{N}}_2 as the dominant atmosphere, and the flue gas is almost entirely composed of {\mathrm{C}\mathrm{O}}_2 and {\mathrm{H}}_2 \mathrm{O}, which is beneficial for carbon capture after water vapor condensation.

The advantages of oxy-combustion are as follows: ① The tail flue gas has a {\mathrm{C}\mathrm{O}}_2 concentration of about 95 \mathrm{\%}, allowing {\mathrm{C}\mathrm{O}}_2 to be captured directly; and ② because the {\mathrm{N}}_2 is separated out, the combustion process produces mainly {\mathrm{N}\mathrm{O}}_x based on nitrogen originally present in the fuel, which results in less thermal {\mathrm{N}\mathrm{O}}_x generation, reducing the cost of {\mathrm{N}\mathrm{O}}_x removal [5].

Although oxy-combustion is regarded as a technology that can effectively achieve carbon capture, the air separation unit (ASU), {\mathrm{C}\mathrm{O}}_2 compression and purification unit (CPU), and flue gas recirculation (FGR) unit used in oxy-combustion require energy. As a result, the net efficiency of atmospheric-pressure oxy-combustion is 8 \mathrm{\%} - 12 \mathrm{\%} less than that of air combustion without carbon capture. Since carbon sequestration and EOR operate at high pressure, pressurized oxy-combustion (POC) with CCUS has gained interest as a means of improving the net power efficiency by operating the boiler, ASU, FGR, and CPU at elevated pressure. In this way, there is no need to lower the furnace pressure to the atmospheric pressure, so much less energy is wasted in the pressure-adjustment process, compared with atmospheric-pressure oxy-combustion. Furthermore, in POC, moisture condenses at high temperatures, making it easier to recover latent heat and incorporate it into the Rankine cycle, which raises the net efficiency. Babcock Thermo evaluated the performance of POC technology in 2000 [6], and Canada Center for Mineral and Energy Technology (CANMET) and ThermoEnergy examined the economics of POC systems in 2005 [7]. Also in 2005, Instituto Tren-tino Edilizia Abitativa (ITEA) proposed a pressurized flameless boiler technology-ISOTHERM, which is shown in Fig. 2 [8]. In this technology, coal-water slurry is first burnt in a pressurized combustor via flameless combustion. Then, high-temperature flue gas is sent to a convective heat-recovery steam generator (HRSG). The net efficiency increases by 3 \mathrm{\%} as the pressure is increased from 0.1 to 1.0 \mathrm{M} \mathrm{P} \mathrm{a}, but further increases in pressure have diminishing return [9,10]. However, the efficiency at 1.0 \mathrm{M} \mathrm{P} \mathrm{a} is still low, with a value of 33.5 \mathrm{\%}.

Gopan et al. [11] at Washington University in St. Louis, USA, proposed a staged POC (SPOC) technology that controls the furnace temperature by staging combustion in a way that significantly reduces the amount of FGR. A schematic diagram of SPOC is shown in Fig. 3 [12]. The principle behind staging is that a portion of the flue gas from the earlier stage is supplied as the dilution gas for the subsequent stage, thereby minimizing flue gas recycling. A comparison by Hagi et al. [13] concluded that the net efficiency of SPOC is 3.9% higher than that of ISOTHERM. Gopan et al. [14] studied the economics of SPOC and found that the efficiency could be increased by 6 \mathrm{\%} when the pressure is increased from 0.1 to 1.6 \mathrm{M} \mathrm{P} \mathrm{a} and staging is incorporated.

The SPOC technology has many advantages: ① The saturation temperature of the water increases with the pressure such that the latent heat can be recovered, thereby increasing the net plant efficiency; ② the gas volume is greatly decreased at high pressure and the size of the boiler is reduced, lowering the footprint; ③ convective heat transfer is enhanced at elevated pressure; ④ the staging allows for highly flexible operation; and {\mathrm{S}\mathrm{O}}_x and {\mathrm{N}\mathrm{O}}_x can be removed simultaneously and economically in a direct contact column (DCC) [15].

Lab-scale experimental facilities have been built to study the emission characteristics of POC, as shown in Table 1 [12,16⇓⇓⇓⇓⇓⇓⇓⇓-25]. Such facilities include a pressurized fluidized bed (FB) reactor, pressurized fixed-bed reactor, and pressurized entrained flow reactor (EFR). Lab-scale pressurized FBs and pressurized fixed-bed reactors have been employed to study the pollutant emissions in an FB during POC, and pressurized EFRs and drop tube furnaces (DTFs) have been used to study pollutants emissions in a PC boiler during POC. Pressurized DTFs have been built at Xi’an Jiaotong University (China) [16], Harbin Institute of Technology (China) [17], and North China Electric Power University (China) [18], while pressurized fixed beds have been constructed at Zhejiang University (China) [19] and Harbin Institute of Technology (China) [26]. Southeast University (China) constructed pressurized FBs with respective capacities of 10 [20] and 30 kilowatt-thermal \left({\mathrm{k}\mathrm{W}}_{\mathrm{t}\mathrm{h}}\right)[21].

Many institutes have built POC demonstrations to test the performance of POC. Dobó et al. [22] at the University of Utah, USA, built a demonstration 300 {\mathrm{k}\mathrm{W}}_{\text{ th }} pressurized entrained flow combustor in which coal is fed into the furnace as a coal-water slurry; the combustor was successfully operated at 1.4 \mathrm{M} \mathrm{P} \mathrm{a}. Yang et al. [12] at Washington University in St. Louis, USA, constructed a dry-feed 100 {\mathrm{k}\mathrm{W}}_{\text{ th }} POC combustor and operated it at 1.5 \mathrm{M} \mathrm{P} \mathrm{a} with stable combustion at a wide range of feed rates. Gardner et al. [23] at Brigham Young University, USA, designed a POC reactor for firing rates of 100 {\mathrm{k}\mathrm{W}}_{\mathrm{t}\mathrm{h}} at pressures up to 3.00 \mathrm{M} \mathrm{P} \mathrm{a}, which was demonstrated to run successfully at 1.36 \mathrm{M} \mathrm{P} \mathrm{a} and a firing rate of 89 {\mathrm{k}\mathrm{W}}_{\mathrm{t}\mathrm{h}}.

As for industrial application, the only industrial-level POC was constructed by ITEA [24], which built and operated a 5 {\mathrm{M}\mathrm{W}}_{\text{ th }} lab-scale flameless POC (FPO) combustor. ITEA is now planning to build a 25 and 50 MW FPO power plant and eventually put it into commercialization. There is still a significant gap between the theoretical understanding and commercialization of POC. The net efficiency of SPOC is still 4 \mathrm{\%} lower than that of traditional air combustion. Furthermore, the design and operation of POC is much more complicated than those of a traditional air combustion boiler. The main challenges of POC are its system complexity and safety issues. The performance of the devices and their cooperation at high pressure are still unknown. When devices are operated under high pressure, the risk of leakage increases, and high-concentration {\mathrm{C}\mathrm{O}}_2 is a threat to human life. In order to address these issues and move this technology toward commercialization, significant work is required. Fortunately, pressurized gasifiers have been effectively validated at the industrial level, which may offer design and operating experience for POC boilers.

Aside from experimental methods, modeling methods have been widely used to predict the performance of POC. Aspen software has been used to calculate the economics of a POC system and optimize its net efficiency [15], Fluent has been used to simulate the temperature distribution and combustion performance of POC [14], and Chemkin has been employed to study POC reaction kinetics [27]. Some conditions have not been realized in a real POC furnace. As a result, this paper makes use of kinetic and thermodynamic studies to forecast the performance of POC.

The major components of the flue gas in \mathrm{P} \mathrm{O} \mathrm{C} are {\mathrm{C}\mathrm{O}}_2 , {\mathrm{H}}_2 \mathrm{O}, and a small proportion of non-condensable gases such as {\mathrm{O}}_2 , \mathrm{A} \mathrm{r}, and {\mathrm{N}}_2. The residual components consist of acid gases such as sulfur oxides \left({\mathrm{S}\mathrm{O}}_x\right) and nitrous oxides \left({\mathrm{N}\mathrm{O}}_x\right). These, along with oxygen, can cause corrosion problems in the pipelines used for transportation. Thus, when the captured {\mathrm{C}\mathrm{O}}_2 is destined for EOR or carbon sequestration and storage, these gases must be controlled to meet the pipeline purity standards [28]. The limitations of {\mathrm{N}\mathrm{O}}_x and {\mathrm{S}\mathrm{O}}_x concentrations are 100 and 50 parts per million by volume (ppmv), respectively [29,30]. Particulate matter (PM) and mercury emissions and control are additional issues that must be addressed [31].

We have conducted extensive research on the aforementioned major pollutants in POC, investigating {\mathrm{S}\mathrm{O}}_3 formation in POC [32], the catalytic effect of {\mathrm{F}\mathrm{e}}_2 {\mathrm{O}}_3 on {\mathrm{S}\mathrm{O}}_3 formation [33], and in-furnace desulfurization in POC [34]. We have also performed work on {\mathrm{N}\mathrm{O}}_x formation and control in POC [35], including {\mathrm{N}\mathrm{O}}_2 formation [36], pressurized selective non-catalytic reduction (SNCR) [27], {\mathrm{N}\mathrm{O}}_x evolution [37], pressurized reburning [38], and \mathrm{N} \mathrm{O} reduction characteristics [39]. Moreover, we contributed to the investigation of a PM sampling method and PM formation characteristics in POC [40⇓-42]. We have built experimental facilities, such as a pressurized fixed bed and pressurized DTF.

Based on a long-term study of the key issues above, we write this review on multi-pollutant formation and control in POC, with the aim of enhancing the understanding of pollutant formation and control in POC. This article reviews recent advances regarding the formation and control of {\mathrm{S}\mathrm{O}}_x , {\mathrm{N}\mathrm{O}}_x , \mathrm{P} \mathrm{M}, and mercury in POC. Section 2 provides a detailed overview of sulfur evolution and of {\mathrm{S}\mathrm{O}}_x formation and removal technology in POC. Section 3 focuses on the migration of nitrogen and on {\mathrm{N}\mathrm{O}}_x formation and control methods in POC. Section 4 reviews PM formation and removal methods during POC. Section 5 provides a discussion on mercury formation and control technology. Finally, a summary and outlook on the removal of pollutants during POC is presented in Section 6.

Sulfur is one of the main elements in coal, and its content differs significantly depending on differences in the depositional environments; however, most sulfur content ranges from 0.2 \mathrm{\%} - 11.0 \mathrm{\%}. The formation of sulfur oxides is a major concern in POC, as they can form sulfuric acid and cause corrosion of the pipelines and other equipment. This section reviews {\mathrm{S}\mathrm{O}}_x evolution, formation, and control in POC.

The forms of sulfur in coal are very abundant and complex. According to differences in chemical structure, sulfur is classified into organic sulfur, inorganic sulfur, and elemental sulfur [43]. Inorganic sulfur includes sulfates and sulfides, taking the main form of pyrite \left({\mathrm{F}\mathrm{e}\mathrm{S}}_2\right). Organic sulfur includes all sulfur combined with a hydrocarbon matrix. Elemental sulfur can also appear in coal at a proportion less than 0.2 \mathrm{\%}.

The sulfur evolution route during coal combustion is outlined in Fig. 4 [44]. Sulfur in coal first decomposes and evolves into {\mathrm{H}}_2 \hspace{0.33em} \mathrm{S}, \mathrm{C} \mathrm{O} \mathrm{S} , {\mathrm{S}\mathrm{O}}_2, and {\mathrm{C}\mathrm{S}}_2 in the gas phase, as well as other compounds in the tar and char, which then burn in an oxidizing atmosphere to produce {\mathrm{S}\mathrm{O}}_2 and {\mathrm{S}\mathrm{O}}_3 [45]. {\mathrm{S}\mathrm{O}}_3 can react with {\mathrm{H}}_2 \mathrm{O} to form {\mathrm{H}}_2 {\mathrm{S}\mathrm{O}}_4, which condenses when the temperature falls below the acid’s dew point.

There are many reports about sulfur evolution under an atmospheric-pressure {\mathrm{C}\mathrm{O}}_2 atmosphere. Tian et al. [46] employed thermogravimetry Fourier-transform infrared (TG-FTIR) analysis to study sulfur evolution in oxy-combustion. Their research revealed that pyrite, a form of iron disulfide \left({\mathrm{F}\mathrm{e}\mathrm{S}}_2\right) can react with {\mathrm{H}}_2 , \mathrm{C} \mathrm{O}, and \mathrm{C} to generate {\mathrm{H}}_2 \hspace{0.33em} \mathrm{S} , \mathrm{C} \mathrm{O} \mathrm{S}, and {\mathrm{C}\mathrm{S}}_2. Organic sulfur is first pyrolyzed to produce sulfur radicals and then transformed into {\mathrm{H}}_2 \hspace{0.33em} \mathrm{S} , \mathrm{C} \mathrm{O} \mathrm{S}, and {\mathrm{C}\mathrm{S}}_2. All sulfur compounds are oxidized to {\mathrm{S}\mathrm{O}}_2 and a small proportion of {\mathrm{S}\mathrm{O}}_3. At a high temperature, sulfate decomposes to produce {\mathrm{S}\mathrm{O}}_2. The researchers noted that {\mathrm{H}}_2 \hspace{0.33em} \mathrm{S} and \mathrm{C} \mathrm{O} \mathrm{S} formation are promoted due to the increase in \mathrm{C} \mathrm{O} generated by gasification in a {\mathrm{C}\mathrm{O}}_2 atmosphere, in comparison with a {\mathrm{N}}_2 atmosphere. In addition, a higher concentration of \mathrm{C} \mathrm{O} promotes the reduction of sulfite to produce more {\mathrm{S}\mathrm{O}}_2. The major sulfur-containing compound formed during combustion in an {\mathrm{N}}_2 atmosphere is {\mathrm{H}}_2 \hspace{0.33em} \mathrm{S}, whereas the equivalent sulfur-containing compound in a {\mathrm{C}\mathrm{O}}_2 atmosphere is COS. Duan et al. [47] also found that COS formation is extremely high in a {\mathrm{C}\mathrm{O}}_2 atmosphere in comparison with a {\mathrm{N}}_2 atmosphere, which might be due to the reduction of {\mathrm{F}\mathrm{e}\mathrm{S}}_2 by CO (Eq. (1)). However, Chen and Bhattacharya [48] stated that the quantity of {\mathrm{H}}_2 \hspace{0.33em} \mathrm{S} is greater than that of COS at 1073 \hspace{0.33em} \mathrm{K} in a {\mathrm{C}\mathrm{O}}_2 atmosphere. When the temperature exceeds 1173 \hspace{0.33em} \mathrm{K} , {\mathrm{C}\mathrm{S}}_2 becomes the predominant sulfur gas.

There are few reports on sulfur evolution in a pressurized pyrolysis process. Liang et al. [19] explored sulfur evolution in a pressurized fixed-bed reactor under pressurized pyrolysis conditions. They collected the pyrolysis gas and analyzed it in a gas chromatography device. As shown in Fig. 5 [19], the researchers noticed that COS and {\mathrm{H}}_2 \hspace{0.33em} \mathrm{S} were the dominant sulfur-containing gases, and {\mathrm{H}}_2 \hspace{0.33em} \mathrm{S} formation was higher than that of COS. As the pressure increased, {\mathrm{H}}_2 \hspace{0.33em} \mathrm{S} decreased significantly, while COS increased first and then decreased at a pressure above 0.6 \mathrm{M} \mathrm{P} \mathrm{a}. The researchers’ primary explanation was that the sulfur self-retention reaction was enhanced at elevated pressure, which was further demonstrated by the increase in the sulfur content in the char.

The evolution of sulfur in POC is complicated and requires further investigation. Liang et al. [49] analyzed the sulfur conversion mechanism in POC; the main pathways are shown in Fig. 6. The sulfur in the fuel (fuel-S) is decomposed to produce COS and {\mathrm{H}}_2 \hspace{0.33em} \mathrm{S}. \mathrm{S} \mathrm{H} is the main intermediate; it reacts with \mathrm{O} to produce \mathrm{S} \mathrm{O}, which is easily oxidized by \mathrm{O} to form {\mathrm{S}\mathrm{O}}_2. HSO is another important precursor for {\mathrm{S}\mathrm{O}}_2 formation. {\mathrm{S}\mathrm{O}}_2 reacts with \mathrm{O} and \mathrm{O} \mathrm{H} to form {\mathrm{S}\mathrm{O}}_3.

In this section, we discuss the formation of {\mathrm{S}\mathrm{O}}_2 and {\mathrm{S}\mathrm{O}}_3 in POC.

{\mathrm{S}\mathrm{O}}_2 emissions in POC have been extensively studied in FB combustion over a temperature range of 850 - {950}^{\circ } \mathrm{C}. Many researchers have observed that {\mathrm{S}\mathrm{O}}_2 emissions are significantly reduced in a FB during POC, as shown in Fig. 7(a) [18,50,51]. On a lab-scale pressurized FB, Lasek et al. [25] discovered that {\mathrm{S}\mathrm{O}}_2 emissions were reduced by more than 30 \mathrm{\%} in POC compared with atmospheric-pressure air combustion. Duan et al. [20] studied {\mathrm{S}\mathrm{O}}_2 emissions in POC on a pressurized FB and noticed that they decreased with increasing pressure. The researchers found that, under pressurized conditions, {\mathrm{S}\mathrm{O}}_2 emissions were significantly lower in an {\mathrm{O}}_2 / {\mathrm{C}\mathrm{O}}_2 atmosphere than in an {\mathrm{O}}_2 / {\mathrm{N}}_2 atmosphere. According to X-ray diffraction (XRD) analysis, the {\mathrm{C}\mathrm{a}\mathrm{S}\mathrm{O}}_4 of ash increases in POC, indicating that POC promotes the ash’s sulfur self-retention effect. As shown in Fig. 7(b), Tang et al. [52] revealed on a pressurized fixed bed that, as the pressure increased, {\mathrm{S}\mathrm{O}}_2 emissions decreased and the ash’s sulfur content increased. When the pressure reached 0.7 \mathrm{M} \mathrm{P} \mathrm{a}, the sulfur in the ash accounted for 90 \mathrm{\%}. Liu et al. [53] studied the {\mathrm{S}\mathrm{O}}_x emissions of coal and biomass co-firing in a 10 {\mathrm{k}\mathrm{W}}_{\mathrm{t}\mathrm{h}} \mathrm{F} \mathrm{B} and noticed that {\mathrm{S}\mathrm{O}}_2 emissions decreased significantly with increasing pressure. The ash’s sulfur content increased at elevated pressure, and the sulfur’s self-retention efficiency increased to nearly double at 0.3 \mathrm{M} \mathrm{P} \mathrm{a} in comparison with 0.1 \mathrm{M} \mathrm{P} \mathrm{a}. Zan et al. [50] investigated {\mathrm{S}\mathrm{O}}_2 emissions in a 100 {\mathrm{k}\mathrm{W}}_{\text{ th }} \mathrm{F} \mathrm{B} and also noticed decreased {\mathrm{S}\mathrm{O}}_2 emissions and enhanced sulfur self-retention in POC. Liang et al. [49] discovered that, as the pressure increased from 0.1 to 1.6 \mathrm{M} \mathrm{P} \mathrm{a}, the {\mathrm{S}\mathrm{O}}_2 emissions decreased by 50 \mathrm{\%} ; the {\mathrm{S}\mathrm{O}}_2 emissions also decreased as the {\mathrm{O}}_2 concentration increased. Wang et al. [54] noticed that the proportion of the fuel sulfur that was converted to {\mathrm{S}\mathrm{O}}_2 decreased at 3 \mathrm{M} \mathrm{P} \mathrm{a}, possibly due to increased {\mathrm{S}\mathrm{O}}_3 conversion under pressurization. Pang et al. [51] measured the {\mathrm{S}\mathrm{O}}_2 emissions in a 30 {\mathrm{k}\mathrm{W}}_{\text{ th }} pressurized FB combustor and discovered that the sulfate content in the ash increased in POC, compared with that in atmospheric-pressure oxy-combustion.

Very few studies discuss PC combustion in POC. Dobó et al. [22] tested {\mathrm{S}\mathrm{O}}_2 emissions in a 300 \hspace{0.33em} \mathrm{k} \mathrm{W} pilot-scale combustor and observed lower {\mathrm{S}\mathrm{O}}_2 emissions in POC between 1300 and {1600}^{\circ } \mathrm{C}. On a pressurized DTF, Lei et al. [18] observed that {\mathrm{S}\mathrm{O}}_2 emissions decreased with increasing pressure and were lower under POC than pressurized air combustion. When the pressure increased to 1.1 \mathrm{M} \mathrm{P} \mathrm{a}, the {\mathrm{S}\mathrm{O}}_2 emissions were less than 50 \mathrm{m} \mathrm{g} \cdot {\mathrm{M}\mathrm{J}}^{-1}, which further decreases {\mathrm{S}\mathrm{O}}_2 emissions. The maximum temperature in that experiment was 1200 {}^{\circ } \mathrm{C}.

The reasons for the decrease in {\mathrm{S}\mathrm{O}}_2 in POC are as follows: ① The partial pressure of {\mathrm{O}}_2 increases as the pressure increases, facilitating the transformation of {\mathrm{S}\mathrm{O}}_2 to {\mathrm{S}\mathrm{O}}_3 ; ② the increase in the partial pressure of {\mathrm{C}\mathrm{O}}_2 promotes the gasification reaction, generating more \mathrm{C} \mathrm{O}, which benefits the reduction of {\mathrm{S}\mathrm{O}}_2 ; ③ sulfur self-retention is enhanced in POC, as it permits {\mathrm{S}\mathrm{O}}_x to react with alkali and alkaline earth metals to form sulfate; ④ the increased pressure extends the contact time of {\mathrm{S}\mathrm{O}}_2 on the char and ash particles, promoting sulfur reduction and capture; and ⑤ the increased pressure promotes {\mathrm{S}\mathrm{O}}_2 adsorption on the ash surface [55].

It should be pointed out that none of these experiments considered the influence of FGR. If FGR is added, {\mathrm{S}\mathrm{O}}_2 emissions may be higher. More research is needed to determine the effect of FGR and residence time on {\mathrm{S}\mathrm{O}}_2 emissions. In addition, the mechanism of sulfur evolution is still unclear, requiring further investigation in the future.

During conventional combustion, most sulfur is transformed into {\mathrm{S}\mathrm{O}}_2, while 0.1 \mathrm{\%} - 1.0 \mathrm{\%} of the sulfur is converted to {\mathrm{S}\mathrm{O}}_3. Although not much {\mathrm{S}\mathrm{O}}_3 is produced, it can readily combine with water to generate sulfuric acid, which poses a risk of corrosion to the pipeline at temperatures below the acid dew point. {\mathrm{O}}_2 concentration, {\mathrm{S}\mathrm{O}}_2 concentration, and active-metal catalysis are the key factors influencing {\mathrm{S}\mathrm{O}}_3 formation [56].

The formation of {\mathrm{S}\mathrm{O}}_3 via homogeneous and heterogeneous routes requires discussion in order to comprehend the formation of {\mathrm{S}\mathrm{O}}_x. Homogeneous {\mathrm{S}\mathrm{O}}_3 formation routes include reactions Eqs. (2)-(4). The main reaction is the combination of {\mathrm{S}\mathrm{O}}_2 and \mathrm{O}, which takes place at high temperatures \left(>1150\hspace{0.33em}\mathrm{K}\right) [57]; the second primary formation route occurs in the presence of water vapor, as \mathrm{O} \mathrm{H} reacts with {\mathrm{S}\mathrm{O}}_2, with {\mathrm{H}\mathrm{O}\mathrm{S}\mathrm{O}}_2 as an intermediate [58]:

In addition, it should be noted that {\mathrm{S}\mathrm{O}}_3 formation can be catalyzed by water wall, fly ash, and vanadium-based catalysts. Fly ash from coal combustion has a catalytic effect on {\mathrm{S}\mathrm{O}}_3 generation that is related to the {\mathrm{F}\mathrm{e}}_2 {\mathrm{O}}_3 content, with the highest catalytic temperature occurring at 700 {}^{\circ } \mathrm{C} [59]. {\mathrm{F}\mathrm{e}}_2 {\mathrm{O}}_3 has a stronger catalysis effect than quartz and aluminum [60].

Many researchers have reported that {\mathrm{S}\mathrm{O}}_3 emissions are elevated under oxy-combustion. Wall et al. [4] determined that the {\mathrm{S}\mathrm{O}}_3 emissions from oxy-combustion were 2.5 - 3.0 times greater than those from air combustion. Croise and Thambimuthu [61] discovered that the conversion ratio of fuel sulfur to {\mathrm{S}\mathrm{O}}_2 changed from 91 \mathrm{\%} in air combustion to 75 \mathrm{\%} in an oxy-combustion atmosphere and 64% in oxy-combustion with FGR. Ahn et al. [62] discovered on a 1.5 \mathrm{M} \mathrm{W} PC furnace that {\mathrm{S}\mathrm{O}}_3 emissions were 4-6 times higher in oxy-combustion than in air combustion when burning high-sulfur coal, while the same {\mathrm{S}\mathrm{O}}_3 emissions in air combustion and oxy-combustion were observed when burning low-sulfur coal. However, in an FB combustion experiment, Roy et al. [63] observed lower {\mathrm{S}\mathrm{O}}_3 emissions in oxy-combustion than in air combustion but found an opposite trend in a DTF.

Some scholars have reported that there is little impact on {\mathrm{S}\mathrm{O}}_3 generation when {\mathrm{C}\mathrm{O}}_2 replaces {\mathrm{N}}_2 as the primary gas. Xiang et al. [64] found that \mathrm{N} \mathrm{O} facilitated {\mathrm{S}\mathrm{O}}_3 formation, and a {\mathrm{C}\mathrm{O}}_2 atmosphere had no effect on {\mathrm{S}\mathrm{O}}_3 generation. Duan et al. [65] studied {\mathrm{S}\mathrm{O}}_3 homogeneous formation in a tube furnace; their results indicated that switching the atmosphere from {\mathrm{O}}_2 / {\mathrm{N}}_2 to {\mathrm{O}}_2 / {\mathrm{C}\mathrm{O}}_2 resulted in negligible changes, and increasing the {\mathrm{H}}_2 \mathrm{O} content promoted the {\mathrm{S}\mathrm{O}}_3 formation rate. Wang et al. [66] investigated {\mathrm{S}\mathrm{O}}_3 formation in a fixed-bed reactor and observed a tiny distinction in {\mathrm{S}\mathrm{O}}_3 formation between an {\mathrm{O}}_2 / {\mathrm{N}}_2 and {\mathrm{O}}_2 / {\mathrm{C}\mathrm{O}}_2 atmosphere. Dai et al. [67] employed density functional theory and revealed that the adsorption energy of {\mathrm{C}\mathrm{O}}_2 on {\mathrm{F}\mathrm{e}}_2 {\mathrm{O}}_3 is smaller than that of {\mathrm{N}}_2, {\mathrm{S}\mathrm{O}}_2, and {\mathrm{O}}_2 ; thus, {\mathrm{C}\mathrm{O}}_2 has minimal impact on {\mathrm{S}\mathrm{O}}_3 generation catalyzed by {\mathrm{F}\mathrm{e}}_2 {\mathrm{O}}_3.

In terms of {\mathrm{S}\mathrm{O}}_3 formation in POC, few experimental measurements have been reported due to the difficulty of {\mathrm{S}\mathrm{O}}_3 sampling and measurement, and the majority of the available literature is based on simulations. Wang et al. [68] calculated {\mathrm{S}\mathrm{O}}_3 formation in POC and stated that, as the pressure increased from 0.1 to 1.5 \mathrm{M} \mathrm{P} \mathrm{a}, the conversion ratio of {\mathrm{S}\mathrm{O}}_2 to {\mathrm{S}\mathrm{O}}_3 could be significantly improved. Wang et al. [32] revealed a strong coupling promotion between {\mathrm{N}\mathrm{O}}_x and {\mathrm{S}\mathrm{O}}_x in \mathrm{P} \mathrm{O} \mathrm{C}, where the existence of {\mathrm{N}\mathrm{O}}_x greatly promotes {\mathrm{S}\mathrm{O}}_3 generation. As depicted in Fig. 8 [68], {\mathrm{S}\mathrm{O}}_3 formation is significantly enhanced with NO at high pressure and low temperature. Since the residence time of the flue gas in POC increases, {\mathrm{S}\mathrm{O}}_3 formation can be significantly promoted. Dai et al. [33] investigated {\mathrm{S}\mathrm{O}}_3 formation catalyzed by {\mathrm{F}\mathrm{e}}_2 {\mathrm{O}}_3 in POC, in what was a rare report on {\mathrm{S}\mathrm{O}}_3 formation in POC. They reported that the catalysis effect of {\mathrm{F}\mathrm{e}}_2 {\mathrm{O}}_3 was weakened at elevated pressure. In the presence of \mathrm{N} \mathrm{O} , {\mathrm{S}\mathrm{O}}_3 formation was promoted.

In conclusion, there are no experimental reports on {\mathrm{S}\mathrm{O}}_3 emissions in POC owing to the difficulty of {\mathrm{S}\mathrm{O}}_3 sampling. Simulations have demonstrated that the conversion ratio of {\mathrm{S}\mathrm{O}}_2 to {\mathrm{S}\mathrm{O}}_3 can be considerably greater in POC than in air combustion. A sample method for {\mathrm{S}\mathrm{O}}_3 under high pressure needs to be developed. Testing and production mechanisms of {\mathrm{S}\mathrm{O}}_3 in POC will be the subject of extensive research in the future.

When compared with atmospheric-pressure oxy-combustion, POC emits less {\mathrm{S}\mathrm{O}}_2 ; however, sulfur capture is still required to decrease the {\mathrm{S}\mathrm{O}}_2 emissions to less than 50 parts per million (ppm). There are two types of sulfur capture technologies: in-furnace sulfur removal and tail gas removal. As with traditional air combustion, limestone can be introduced into the furnace in POC to lower the initial {\mathrm{S}\mathrm{O}}_x pollution. Because the flue gas pressure is high in the tail gas, a DCC can be utilized to capture {\mathrm{S}\mathrm{O}}_x, with a {\mathrm{S}\mathrm{O}}_x -removal efficiency as high as 99 \mathrm{\%}.

The acid dew point of {\mathrm{H}}_2 {\mathrm{S}\mathrm{O}}_4 depends on the {\mathrm{S}\mathrm{O}}_3 concentration and {\mathrm{H}}_2 \mathrm{O} concentration. Therefore, reducing {\mathrm{S}\mathrm{O}}_3 and {\mathrm{H}}_2 \mathrm{O} concentration in the flue gas is useful for decreasing the acid dew point and reducing corrosion. A condensation unit can condense moisture to reduce the {\mathrm{H}}_2 \mathrm{O} content in the flue gas. To decrease {\mathrm{S}\mathrm{O}}_3 formation, in-furnace desulfurization can be applied. For example, the addition of limestone to the furnace can capture sulfur and reduce {\mathrm{S}\mathrm{O}}_3 formation.

With the advantages of a widespread source, easy availability, and low price, limestone is the most traditional desulfurization agent used in circulating FBs (CFBs). The main component of limestone is {\mathrm{C}\mathrm{a}\mathrm{C}\mathrm{O}}_3. Baker [69] proposed the equilibrium curve of {\mathrm{C}\mathrm{a}\mathrm{C}\mathrm{O}}_3 calcination, as shown in Eq. (5):

where P_{\mathrm{e}} \left(0.1\mathrm{M}\mathrm{P}\mathrm{a}\right) is the {\mathrm{C}\mathrm{O}}_2 equilibrium partial pressure and T is the temperature(K).

The calcination equilibrium curve is drawn in Fig. 9 [34], which shows that the temperature and {\mathrm{C}\mathrm{O}}_2 partial pressure can significantly affect the thermal decomposition of {\mathrm{C}\mathrm{a}\mathrm{C}\mathrm{O}}_3. For example, when the {\mathrm{C}\mathrm{O}}_2 partial pressure increases to above 0.2 \mathrm{M} \mathrm{P} \mathrm{a} at a conventional desulfurization temperature \left(850-{950}^{\circ }\mathrm{C}\right) , {\mathrm{C}\mathrm{a}\mathrm{C}\mathrm{O}}_3 cannot decompose. According to whether or not {\mathrm{C}\mathrm{a}\mathrm{C}\mathrm{O}}_3 can be decomposed, the sulfation process can be classified into indirect sulfation or direct sulfation. During indirect sulfation, {\mathrm{C}\mathrm{a}\mathrm{C}\mathrm{O}}_3 is first decomposed to \mathrm{C} \mathrm{a} \mathrm{O}, which reacts with {\mathrm{S}\mathrm{O}}_2 and {\mathrm{O}}_2 to produce {\mathrm{C}\mathrm{a}\mathrm{S}\mathrm{O}}_4 (Eq. (7)). When the partial pressure of {\mathrm{C}\mathrm{O}}_2 is greater than the equilibrium pressure of {\mathrm{C}\mathrm{a}\mathrm{C}\mathrm{O}}_3 decomposition, {\mathrm{C}\mathrm{a}\mathrm{C}\mathrm{O}}_3 directly interacts with {\mathrm{S}\mathrm{O}}_2 and {\mathrm{O}}_2 to produce {\mathrm{C}\mathrm{a}\mathrm{S}\mathrm{O}}_4 (Eq. (8)), which is direct sulfation. When the {\mathrm{C}\mathrm{O}}_2 pressure is higher than 0.4 \mathrm{M} \mathrm{P} \mathrm{a} at temperature below 1000 {}^{\circ } \mathrm{C}, the sulfation process changes from indirect sulfation to direct sulfation [70]. Therefore, direct sulfation occurs inPOC.

The main desulfurization processes are as follows [71]:

The decomposition of calcium carbonate

Indirect sulfation

Direct sulfation

The sulfation of limestone can be controlled by kinetics and diffusion. The desulfurization efficiency of limestone is influenced by competition between the sulfation’s the intrinsic rate and the {\mathrm{C}\mathrm{a}\mathrm{S}\mathrm{O}}_4 diffusion rate. When limestone is injected into the high-temperature zone, it decomposes immediately and generates high-reactivity \mathrm{C} \mathrm{a} \mathrm{O}. The reaction rate at this stage in affected by the intrinsic rate. However, due to its large molar volume, {\mathrm{C}\mathrm{a}\mathrm{S}\mathrm{O}}_4 can block small pores and encapsulate {\mathrm{C}\mathrm{a}\mathrm{C}\mathrm{O}}_3, preventing the reaction and reducing the desulfurization efficiency. Therefore, when a {\mathrm{C}\mathrm{a}\mathrm{S}\mathrm{O}}_4 layer is formed, the desulfurization efficiency is influenced by this {\mathrm{C}\mathrm{a}\mathrm{S}\mathrm{O}}_4 diffusion rate.

The desulfurization efficiency is also affected by reaction time. In the initial stages, indirect sulfation desulfurizes more quickly than the direct sulfation process. However, as the reaction progresses, direct sulfation has better desulfurization efficiency than the indirect sulfation process [72]. The \mathrm{C} \mathrm{a} \mathrm{O} formed during the initial stages of the indirect sulfation process has a large specific surface area and high reactivity with {\mathrm{S}\mathrm{O}}_2 and {\mathrm{O}}_2, resulting in a higher desulfurization efficiency [73]. The direct sulfation process generates {\mathrm{C}\mathrm{O}}_2 on the surface of {\mathrm{C}\mathrm{a}\mathrm{C}\mathrm{O}}_3, forming a porous structure in the particle, which decreases {\mathrm{S}\mathrm{O}}_2 ’s diffusion resistance to unre-acted limestone and increases the sulfation efficiency [74].

Many scholars have conducted experimental studies on limestone desulfurization in oxy-combustion, finding that the desulfurization efficiency and optimum desulfurization temperature of oxy-combustion are higher than those of air combustion. Obras-Loscertales et al. [75] studied limestone desulfurization in oxy-combustion using a FB reactor and reported that the desulfurization efficiency of indirect sulfation was greater than that of direct sulfation. The maximum desulfurization efficiency under {\mathrm{O}}_2 / {\mathrm{C}\mathrm{O}}_2 occurs at 900 - 925 {}^{\circ } \mathrm{C}, whereas the maximum desulfurization efficiency under {\mathrm{O}}_2 / {\mathrm{N}}_2 = 35 / 65 occurs at {860}^{\circ } \mathrm{C}. Gómez et al. [76] studied limestone desulfurization in oxy-combustion on a 30 {\mathrm{M}\mathrm{W}}_{\text{ th }} CFB power plant and observed that the desulfurization efficiency of oxy-combustion was greater than that of air combustion, owing to the longer contact time between {\mathrm{S}\mathrm{O}}_2 and the calcium-based absorber in oxy-combustion with FGR. Rahiala et al. [77] found that an improvement in {\mathrm{C}\mathrm{O}}_2 concentration improved the sulfation rate on a pilot-scale CFB. Kim et al. [78] studied direct sulfation in oxy-combustion and applied a shrinking core model to simulate the direct sulfation process.

Researchers have reported an increase in desulfurization efficiency in POC. Pang et al. [51] examined desulfurization in a 30 {\mathrm{k}\mathrm{W}}_{\text{ th }} pressurized FB combustor and found that the desulfurization efficiency improved in POC in comparison with oxy-combustion at atmospheric pressure. Boskovic et al. [79] investigated the desulfurization efficiency on a pressurized CFB and noticed that the desulfurization efficiency increased with an increase in pressure. Qiu and Lindqvist [80] observed that increasing the {\mathrm{C}\mathrm{O}}_2 concentration decreased the sulfation rate when the {\mathrm{S}\mathrm{O}}_2 concentration was kept constant. Dai et al. [34] examined the desulfurization efficiency of limestone in a pressurized DTF. Their main results, shown in Fig. 10 [34], indicated that the desulfurization efficiency increases with increasing pressure, and the optimal desulfurization temperature is higher than that of air combustion.

Other factors that can affect desulfurization efficiency include particle size, water vapor, additives, and so forth. The particle size of the limestone used has a remarkable impact on desulfurization efficiency. On the one hand, because the specific surface area increases with a decrease in particle size, the desulfurization efficiency of small limestone particles is often higher than that of larger particles [81]. On the other hand, fine particles cannot be captured by the cyclone separator in a CFB, resulting in insufficient residence time and lowering the desulfurization efficiency [82]. {\mathrm{H}}_2 \mathrm{O} was found to have a positive effect on desulfurization by elevating solid-state ion diffusion in the {\mathrm{C}\mathrm{a}\mathrm{S}\mathrm{O}}_4 product layer [83]. Additives can promote the desulfurization efficiency of limestone, mainly by making the lattice defective [84].

In conclusion, the following reasons contribute to an increase in desulfurization efficiency in POC system: ① Direct sulfation occurs in POC, generating pores on the surface of the limestone particles and thereby increasing the surface area and decreasing diffusion resistance; ② the partial pressure of {\mathrm{O}}_2 and {\mathrm{S}\mathrm{O}}_2 increases in POC, promoting the sulfation rate; and ③ the contact time of {\mathrm{S}\mathrm{O}}_2 on the ash particles in POC is longer than that in atmospheric-pressure air combustion, leading to a higher capture efficiency. Limestone exhibits high desulfurization efficiency in POC. However, more research is still needed to identify the optimum desulfurization temperature and \mathrm{C} \mathrm{a} / \mathrm{S} ratio.

In conventional combustion, wet flue gas desulfurization (WFGD) is employed to control {\mathrm{S}\mathrm{O}}_2 emissions. The performance of WFGD in a 1 gigawatt-electric \left({\mathrm{G}\mathrm{W}}_{\mathrm{e}}\right) oxy-combustion power plant was evaluated by Neveux et al. [85], whose simulation indicated high desulfurization efficiency in oxy-combustion for both high- and low-sulfur coal. However, in order to reduce cost, they recommended installing WFGD + DCC in an oxy-combustion power plant.

In POC, WFGD could be substituted by a DCC. Due to the increase in the {\mathrm{O}}_2 partial pressure, NO oxidation is accelerated; as a result, {\mathrm{N}\mathrm{O}}_x elimination is improved because {\mathrm{N}\mathrm{O}}_2 dissolves more easily in water than \mathrm{N} \mathrm{O} does. {\mathrm{S}\mathrm{O}}_2 and {\mathrm{S}\mathrm{O}}_3 can be dissolved in water. The interaction between sulfur and nitrogen species further enhances the dissolution of {\mathrm{S}\mathrm{O}}_x and {\mathrm{N}\mathrm{O}}_x. Therefore, DCC is capable of removing the {\mathrm{S}\mathrm{O}}_x and {\mathrm{N}\mathrm{O}}_x in water together. A DCC can be used for both the elimination of {\mathrm{S}\mathrm{O}}_x and {\mathrm{N}\mathrm{O}}_x and the recovery of water’s latent heat. As shown in Fig. 11(a) [15], pressurized flue gas enters the bottom of the DCC, and water is sprayed from the top. {\mathrm{S}\mathrm{O}}_x and {\mathrm{N}\mathrm{O}}_x react with the water and are removed. A DCC has a higher removal efficiency when operated at elevated pressure. In comparison with traditional selective catalytic reduction (SCR) and WFGD, the investment can be decreased by almost70% [86].

The reaction mechanisms of a DCC are shown in Fig. 11(b) [87] and Table 2 [13]. The main reaction pathways are as follows: ① NO oxidation in the gas phase to water-soluble {\mathrm{N}\mathrm{O}}_x ; ② {\mathrm{N}\mathrm{O}}_x dissolution in water; ③ {\mathrm{S}\mathrm{O}}_2 and {\mathrm{S}\mathrm{O}}_3 dissolution in water; and ④ a liquid-phase reaction between {\mathrm{N}\mathrm{O}}_x and {\mathrm{S}\mathrm{O}}_x.

Air Products [88], an industrial gas supplier, conducted a pressurized sour compression experiment, finding that the {\mathrm{N}\mathrm{O}}_{x^{-}} removal efficiency reached 80 \mathrm{\%} when the pressure increased to 1.5 \mathrm{M} \mathrm{P} \mathrm{a}, and that {\mathrm{S}\mathrm{O}}_x emissions could be reduced by more than 99 \mathrm{\%}. In addition, increasing the {\mathrm{S}\mathrm{O}}_2 / \mathrm{N} \mathrm{O} ratio decreased the {\mathrm{S}\mathrm{O}}_2 removal efficiency. Gopan et al. [14] suggested that a DCC could be used in SPOC systems to scrub {\mathrm{S}\mathrm{O}}_x and {\mathrm{N}\mathrm{O}}_x while recovering latent heat. They also reported that a higher \mathrm{S} / \mathrm{N} ratio was favorable for {\mathrm{S}\mathrm{O}}_x and {\mathrm{N}\mathrm{O}}_x removal. Stokie et al. [87] designed a countercurrent packed-bed column DCC for a 100 {\mathrm{k}\mathrm{W}}_{\text{ th }} POC combustor. Their research indicated that it is possible to achieve {\mathrm{S}\mathrm{O}}_{x^{-}} removal efficiencies of up to 100 \mathrm{\%}, as illustrated in Fig. 12 [87]. However, with a residence time of 95 \hspace{0.33em} \mathrm{s} and a pressure of 1.1 \mathrm{M} \mathrm{P} \mathrm{a} , {\mathrm{N}\mathrm{O}}_x -removal efficiency only reached 80 \mathrm{\%}. The {\mathrm{N}\mathrm{O}}_{x^{-}} removal efficiency increased slightly when the {\mathrm{O}}_2 partial pressure increased above 0.1 \mathrm{M} \mathrm{P} \mathrm{a}. Therefore, the researchers concluded that the {\mathrm{O}}_2 concentration in the tail gas can be lower than 1 \mathrm{\%}. The {\mathrm{N}\mathrm{O}}_{x^{-}} removal efficiency declined by 10 \mathrm{\%} as the temperature increased from 24 to {202}^{\circ } \mathrm{C}. Moreover, the \mathrm{N} / \mathrm{S} ratio was found to have a significant impact on the removal efficiency, with low {\mathrm{N}\mathrm{O}}_x concentrations dramatically reducing {\mathrm{S}\mathrm{O}}_x -removal efficiency. The researchers concluded that it is best to maintain a high {\mathrm{N}\mathrm{O}}_x concentration without introducing further denitrification techniques before the DCC.

Because the {\mathrm{N}\mathrm{O}}_x -removal efficiency decreases with increasing temperature, Verma et al. [89] proposed a novel DCC with numerous water inlets. Simulation showed that the optimal scheme was to divide the water into two streams. As indicated in Fig. 13 [89], the flue gas is injected from the bottom, and the water is sprayed from two distinct inlets at different heights. The optimum scrubbing efficiency occurs when 76.0 \mathrm{\%} of the water is fed from the bottom stage and 23.5 \mathrm{\%} is injected from the top. The {\mathrm{S}\mathrm{O}}_{x^{-}} and {\mathrm{N}\mathrm{O}}_{x^{-}} removal efficiency of a DCC with two water inlets is substantially higher than that of a conventional DCC; the two water inlets permit a quick decrease in temperature at the DCC’s bottom and accelerate NO oxidation at low temperatures. The longer residence time and reduced gas volume at a lower temperature are also beneficial for removing {\mathrm{N}\mathrm{O}}_x.

To summarize, a DCC can effectively eliminate {\mathrm{N}\mathrm{O}}_x and {\mathrm{S}\mathrm{O}}_x at low cost and with high efficiency in POC. The disadvantage of a DCC is that its removal efficiency heavily depends on the {\mathrm{N}\mathrm{O}}_x / {\mathrm{S}\mathrm{O}}_2 ratio. However, {\mathrm{N}\mathrm{O}}_x / {\mathrm{S}\mathrm{O}}_2 ratio in POC maybe not the optimum value, which can lead to a low capture efficiency of {\mathrm{S}\mathrm{O}}_x and {\mathrm{N}\mathrm{O}}_x. Thus, it remains important to investigate the mechanisms and improve the removal efficiency of {\mathrm{S}\mathrm{O}}_x and {\mathrm{N}\mathrm{O}}_x in a DCC.

Controlling {\mathrm{N}\mathrm{O}}_x emissions is an important challenge in POC due to the harmful effects of such emissions on equipment, transportation lines, acid rain, environment, and human health. The primary source of {\mathrm{N}\mathrm{O}}_x formation during solid fuel combustion is the oxidation of the nitrogen in the fuel (fuel-N), which is responsible for the majority of {\mathrm{N}\mathrm{O}}_x emissions from the nitrogen species present in the fuel. A smaller proportion of {\mathrm{N}\mathrm{O}}_x produces through two mechanisms: the thermal- {\mathrm{N}\mathrm{O}}_x mechanism, which accounts for 5 \mathrm{\%} - 25 \mathrm{\%} of {\mathrm{N}\mathrm{O}}_x emissions; and the prompt- {\mathrm{N}\mathrm{O}}_x mechanism, which is responsible for less than 5% of emissions [90]. Typically, the generation of thermal {\mathrm{N}\mathrm{O}}_x is considered negligible when temperatures are below 1800 \hspace{0.33em} \mathrm{K}. Conversely, {\mathrm{N}\mathrm{O}}_x produced by the reaction of hydrocarbon radicals and molecular nitrogen is referred to as prompt {\mathrm{N}\mathrm{O}}_x [91], which is characterized by its short duration and increased reactivity under fuel-rich conditions, contributing to fewer than 50 \mathrm{p} \mathrm{p} \mathrm{m} of {\mathrm{N}\mathrm{O}}_x emissions. The fuel- \mathrm{N} mechanism generates {\mathrm{N}\mathrm{O}}_x emissions in the range of 200 - 800 \mathrm{p} \mathrm{p} \mathrm{m}, with the oxidation of gaseous and the residual nitrogen in the char (char- \mathrm{N}) being a major contributor to {\mathrm{N}\mathrm{O}}_x emissions via the fuel- \mathrm{N} mechanism [92].

Fig. 14 provides a simplified diagram modified from Refs. [93- 95 ] that illustrates {\mathrm{N}\mathrm{O}}_x production from fuel-N. The allocation of nitrogen within char and volatiles is predominantly influenced by the type of fuel. At the initial stage of devolatilization, the nitrogen present in the coal is liberated in the form of tar (tar-N), along with a tiny quantity of {\mathrm{N}\mathrm{H}}_3. This tar- \mathrm{N} then undergoes secondary pyrolysis at high temperatures to form \mathrm{H} \mathrm{C} \mathrm{N} and other light gases. The transformation of \mathrm{H} \mathrm{C} \mathrm{N} and ammonia to \mathrm{N} \mathrm{O} or {\mathrm{N}}_2 then occurs, based on the concentration of fuel-N and stoichiometry. At elevated temperature, char-N undergoes heterogeneous conversion to \mathrm{N} \mathrm{O} and {\mathrm{N}}_2 or may release HCN. Loffler et al. [96] suggested that \mathrm{H} \mathrm{C} \mathrm{N} is primarily converted to {\mathrm{N}}_2 \mathrm{O}, while {\mathrm{N}\mathrm{H}}_3 transforms to either \mathrm{N} \mathrm{O} or {\mathrm{N}}_2.

It has been found that pyrolytic processes-including the distribution of nitrogen in the coal (coal-N) into nitrogen in the volatile (volatile-N) and char-N during the initial pyrolysis step-have an impact on the complete process of solid-fuel combustion. While volatile nitrogen is released during the first step of pyrolysis, tar nitrogen is released from the heavy structured compound known as tar during secondary pyrolysis, and char nitrogen remains in the char after devolatilization [97].

Fractionating coal into volatiles and char is vital during coal combustion, as devolatilization produces volatile products more quickly than char [98]. To develop an effective approach for reducing the overall emissions from combustion, it is crucial to comprehend how nitrogen is distributed during devolatilization, as this distribution has a strong influence on pollutant generation. During devolatilization, the most prominent gaseous nitrogen molecules are {\mathrm{N}\mathrm{H}}_3 , \mathrm{H} \mathrm{C} \mathrm{N} , \mathrm{N} \mathrm{O} , {\mathrm{N}}_2 \mathrm{O}, and {\mathrm{N}}_2. During pyrolysis, there is a decrease in the release of volatiles as the pressure increases. This decrease can be attributed to the physical inhibition of devolatilization, which results from an increase in flow resistance caused by the heightened pressure [99].

The main sources of \mathrm{N} \mathrm{O} are the oxidation of {\mathrm{N}\mathrm{H}}_3 and \mathrm{H} \mathrm{C} \mathrm{N}, as well as char-N. These {\mathrm{N}\mathrm{O}}_x precursors \left({\mathrm{N}\mathrm{H}}_3\right. and \left.\mathrm{H}\mathrm{C}\mathrm{N}\right) can be generated from coal-N directly or from tar-N after the secondary reactions [100]. The nitrogen in low-grade coals such as lignite is predominantly present in amine functional groups, which undergo conversion into {\mathrm{N}\mathrm{H}}_3 at relatively low temperatures. At elevated temperatures, superior-grade coal typically contains pyridine, pyrrole, and quaternary nitrogen, formed through the process of pyrolysis. HCN is typically generated through the transformation of pyridines and pyrrole nitrogen. {\mathrm{C}\mathrm{O}}_2 affects the production of {\mathrm{N}\mathrm{H}}_3 and \mathrm{H} \mathrm{C} \mathrm{N}, as {\mathrm{C}\mathrm{O}}_2 is adsorbed onto the coal surface matrix at comparatively lower temperatures and promotes the generation of \mathrm{H} \mathrm{C} \mathrm{N} at higher pressures due to the active \mathrm{C} / {\mathrm{C}\mathrm{O}}_2 bond. Duan et al. [94] conducted a series of experiments and found that the concentration of HCN increased more prominently in a {\mathrm{C}\mathrm{O}}_2 environment than in {\mathrm{N}}_2 as the pressure increased from 0.1 to 0.7 \mathrm{M} \mathrm{P} \mathrm{a}, as illustrated in Fig. 15(a). A minor rise in {\mathrm{N}\mathrm{H}}_3 concentration was observed as the pressure increased, as illustrated in Fig. 15(b) [101], which was attributed to the hydrogenation of char-N by H-radicals.

When coal is devolatilized, the nitrogen present in the coal evolves along with the tar. At elevated temperature, tar undergoes a transformation to produce \mathrm{H} \mathrm{C} \mathrm{N} and other light gases. During this process, two types of volatile- \mathrm{N} are produced: rapid volatile- \mathrm{N}, which is found in light tar materials that produce {\mathrm{N}\mathrm{H}}_3 through pyrolysis reactions; and gradual volatile-N, which is found in heavier tars and predominantly leads to the creation of HCN [102]. This \mathrm{H} \mathrm{C} \mathrm{N} is later transformed into \mathrm{N} \mathrm{O} or {\mathrm{N}}_2, based on the concentration of fuel-N and the local stoichiometry. Therefore, at a higher pressure, tar undergoes more secondary exothermic reactions than it does at atmospheric pressure. These chemical reactions may include either the breakdown of molecules, resulting in the formation of smaller hydrocarbons, or the rebuilding of molecules, resulting in the creation of char [103,104].

Elevating the pressure causes a reduction in the amount of tar and an increase in char and gas formation, since the primary product and char take a longer time to interact under high-pressure conditions. According to research conducted by Cai et al. [105], a higher heating rate enhanced the net nitrogen transfer into tar. However, an increase in pressure impeded this transfer. The researchers concluded that the distribution of nitrogen in coal during pyrolysis is heavily dependent on the type of coal being used. With some types of coal, an increase in pressure results in a significant decrease in the amount of nitrogen present in the volatile and tar phases. However, for other types of coal, the effect of pressure on nitrogen distribution is minimal.

High-quality coal tars produce char at elevated temperatures instead of breaking down into simpler gaseous compounds, whereas tars from biomass and peat do not exhibit this behavior [104]. In Loy Yang brown coal pyrolysis in an argon atmosphere, increasing the pressure increases the {\mathrm{N}\mathrm{H}}_3 yield significantly but only minimally affects the \mathrm{H} \mathrm{C} \mathrm{N} yield, resulting in increased char and {\mathrm{N}\mathrm{H}}_3 production, as shown in Fig. 16 [106]. As the pressure increases, volatile precursors are retained for an extended period inside the char, resulting in an increased concentration of the activate site on the surface of the char. This, in turn, causes greater generation of soot when the char and volatiles interact, ultimately increasing char production. Laughlin et al. [107] found that pressure did not affect the ratio of \mathrm{N} to \mathrm{C} in char during coal pyrolysis in a {\mathrm{N}}_2 atmosphere. Maliutina et al. [108] studied biomass pyrolysis in a {\mathrm{N}}_2 atmosphere and noted that, when the pressure rose to 1.5 \mathrm{M} \mathrm{P} \mathrm{a}, there was an increase in nitrogen concentration in biomass char; however, beyond 1.5 \mathrm{M} \mathrm{P} \mathrm{a}, the concentration decreased. That said, there is no agreement on how pressure affects the distribution of char-N.

Increasing the operating pressure can reduce the concentrations of {\mathrm{N}\mathrm{O}}_x in combustion, as shown in Fig. 17 [20,21,109⇓-111]. The observed enhanced decrease in NO emissions may be attributed to the increased partial pressure of {\mathrm{O}}_2 and the consequent oxidation of \mathrm{N} \mathrm{O} to {\mathrm{N}\mathrm{O}}_2. In addition, the retention of nitrogen-containing species time has been found to be extended within the char particles, enabling \mathrm{N} \mathrm{O} to undergo reduction to {\mathrm{N}}_2 [112]. At increased pressure, the combustion of char takes place at an earlier stage, with a substantial portion of fuel-N transforming into {\mathrm{N}}_2 due to the low concentration of oxygen [113]. Furthermore, as the pressure increases, the diffusion coefficient of {\mathrm{O}}_2 to the char surface declines, which in turn impedes the motion of {\mathrm{O}}_2 to the char, contributing to low {\mathrm{N}\mathrm{O}}_x formation resulting from fuel-N [21,114].

Several researchers have explained the NO reduction mechanism: The increase in the partial pressure of {\mathrm{C}\mathrm{O}}_2 provides a good environment for coal gasification, which results in the formation of CO through reaction (Eq. (14)). In addition, {\mathrm{C}\mathrm{O}}_2 can decompose/- gasify into CO through reaction (Eq. (15)).

and

CO reduces NO homogeneously via reaction (Eq. (16)), and a higher pressure increases the NO retention time within the char pores, causing heterogeneous NO reduction (Eq. (17)) [115]. Therefore, the decrease in NO emissions with pressure is attributed to the favorable conditions for reactions (Eqs. and (17)).

Fig. 18 [20,109,110] illustrates the conclusions of different studies on {\mathrm{N}}_2 \mathrm{O} emissions under high pressure that reported inconsistent results: Svoboda and Pohořelý [109] reported a slight increase in {\mathrm{N}}_2 \mathrm{O} emissions with increasing pressure, whereas \mathrm{L} \mathrm{u} et al. [113] found a decreasing pattern similar to that of NO emissions. The reasons for this pressure effect on {\mathrm{N}}_2 \mathrm{O} evolution are not yet clear, and there is no definitive explanation for it. Some studies have suggested that {\mathrm{N}}_2 \mathrm{O} emissions increase due to {\mathrm{N}\mathrm{O}}_x reduction, while others have proposed that {\mathrm{N}}_2 \mathrm{O} emissions decrease with pressure, similar to NO emissions. One study [116] suggested that the combustion environment affects {\mathrm{N}}_2 \mathrm{O} emissions, while other studies have found little effect. In addition, the bed material in pressurized FB combustion (PFBC) can act as a catalyst or adsorbent for {\mathrm{N}\mathrm{O}}_x formation through heterogeneous reactions, which increases the difficulty of understanding PFBC. Therefore, more research is required to elaborate the mechanism behind {\mathrm{N}}_2 \mathrm{O} evolution in pressurized coal combustion.

The influence of pressure on \mathrm{N} \mathrm{O} and {\mathrm{N}}_2 \mathrm{O} during oxy- and air-fired combustion at elevated pressure was explored by Lasek et al. [25], who found that an increase in pressure had a stronger influence on suppressing the generation of \mathrm{N} \mathrm{O} and {\mathrm{N}}_2 \mathrm{O} under air-fired conditions, as opposed to oxy-fired conditions. At an elevated pressure, air combustion displays a greater tendency to form {\mathrm{N}\mathrm{O}}_x than oxy-combustion, which can be attributed to the higher concentration of {\mathrm{C}\mathrm{O}}_2 in an oxy-combustion environment in comparison with an air environment. In the Boudouard reaction \left({\mathrm{C}\mathrm{O}}_2\right. + {\mathrm{C}}_{\mathrm{s}} \rightleftharpoons 2 \mathrm{C} \mathrm{O}), {\mathrm{C}\mathrm{O}}_2 reacts with char to form \mathrm{C} \mathrm{O}, which can then react with \mathrm{N} \mathrm{O} in the presence of char (Eq. (16)) and reduce it to {\mathrm{N}}_2.

Duan et al. [20] found that {\mathrm{N}\mathrm{O}}_x emissions (i.e., \mathrm{N} \mathrm{O} and {\mathrm{N}}_2 \mathrm{O} emissions) are significantly lowered at elevated pressure, with a greater reduction occurring in oxy-combustion due to the high {\mathrm{C}\mathrm{O}}_2 concentration promoting carbon gasification and facilitating NO reduction. The homogeneous generation of NO decreases as the pressure increases, as shown by Aho et al. [117], who explained that, at high pressure, the three-body reaction \mathrm{H} + {\mathrm{O}}_2 \left(+\mathrm{M}\right) \rightleftharpoons {\mathrm{H}\mathrm{O}}_2 \left(+\mathrm{M}\right) dominates over \mathrm{H} + {\mathrm{O}}_2 \rightleftharpoons \mathrm{O} \mathrm{H} + \mathrm{O}, causing the \mathrm{O} and \mathrm{O} \mathrm{H} concentration to decrease and resulting in the suppression of {\mathrm{N}\mathrm{O}}_x formation. Therefore, pressurized combustion has the potential to decrease {\mathrm{N}\mathrm{O}}_x and \mathrm{C} \mathrm{O} emissions.

To summarize, the impact of pressure on {\mathrm{N}\mathrm{O}}_x emissions in {\mathrm{O}}_2 / {\mathrm{C}\mathrm{O}}_2 combustion involves three main aspects. Firstly, there is an increased transformation of fuel-N into \mathrm{H} \mathrm{C} \mathrm{N} and {\mathrm{N}\mathrm{H}}_3 under high pressure. Secondly, the retention time of {\mathrm{N}\mathrm{O}}_x in the char pores is extended, resulting in a decrease in {\mathrm{N}\mathrm{O}}_x emissions. Lastly, the presence of active radicals-including \mathrm{O} and \mathrm{O} \mathrm{H} -is reduced, inhibiting \mathrm{N} \mathrm{O} formation. At high pressures, the reduction of {\mathrm{N}\mathrm{O}}_x and the oxidation of fuel-N occur together, resulting in a decline in {\mathrm{N}\mathrm{O}}_x emissions. The effect of pressure on decreasing {\mathrm{N}\mathrm{O}}_x emissions is more pronounced than its impact on the formation of {\mathrm{N}\mathrm{O}}_x.

A thorough comprehension of the oxidation of volatile-N in POC is necessary in order to develop a combustion system with low {\mathrm{N}\mathrm{O}}_x emissions [118]. A prominent source of {\mathrm{N}\mathrm{O}}_x is volatile- \mathrm{N} oxidation, which can result in the formation of \mathrm{H} \mathrm{C} \mathrm{N} during the devolatilization of high-grade fuel and coal, or of {\mathrm{N}\mathrm{H}}_3 during the devolatilization of low-quality coal and biomass. Under certain reaction conditions, \mathrm{H} \mathrm{C} \mathrm{N} and {\mathrm{N}\mathrm{H}}_3 oxidation lead to the generation of \mathrm{N} \mathrm{O} ; under other conditions, they are reduced to {\mathrm{N}}_2.

Although recent studies have investigated the reaction chemistry of HCN at ambient pressure, the oxidation mechanism of HCN and its related chemistry under elevated pressure and temperature remain insufficiently investigated. To understand how the oxidation of HCN occurs at elevated pressures, we have used the Glarborg et al. [119] mechanism to predict the behavior of HCN oxidation, according to Chemkin simulations (Fig. 19). It was observed that the oxidation temperature of HCN increases with an increase in pressure. However, the selectivity toward NO and {\mathrm{N}}_2 \mathrm{O} decreases with increasing pressure. At higher pressures, there is a decrease in the amounts of \mathrm{N} \mathrm{O} and {\mathrm{N}}_2 \mathrm{O} emissions. The remaining \mathrm{N} -containing components are {\mathrm{N}}_2 and a small proportion of {\mathrm{N}\mathrm{O}}_2. The impact of pressure on these chemical reactions is significant when the pressure is below 1 \mathrm{M} \mathrm{P} \mathrm{a}, with minor variations being observed above that pressure [120]. While HCN oxidation is well understood at atmospheric pressure, more research is needed to comprehend its oxidation chemistry under elevated pressure. The rate of production analysis indicates that certain reactions are particularly important for the consumption of \mathrm{H} \mathrm{C} \mathrm{N} under pressurized conditions:

During the early stages of coal [121] or biomass [122] combustion and/or gasification, {\mathrm{N}\mathrm{H}}_3 is generated. The SNCR process predominantly employs {\mathrm{N}\mathrm{H}}_3 to reduce {\mathrm{N}\mathrm{O}}_x ; researchers have also explored the potential of {\mathrm{N}\mathrm{H}}_3 as an energy carrier devoid of carbon. The oxidation of solid fuels during combustion leads to the formation of \mathrm{N} \mathrm{O}, and {\mathrm{N}\mathrm{H}}_3 is a significant contributor to this process. However, the conversion of {\mathrm{N}\mathrm{H}}_3 to either \mathrm{N} \mathrm{O} or {\mathrm{N}}_2 depends on various reaction parameters. The conversion of volatile-N to NO or {\mathrm{N}}_2 is mainly determined by the amine radical structure, such as \mathrm{N} , \mathrm{N} \mathrm{H}, or {\mathrm{N}\mathrm{H}}_2. Various investigations have conducted modeling of ammonia chemistry during combustion, but there is a lack of research on the oxidation of {\mathrm{N}\mathrm{H}}_3 at elevated pressure.

To address this gap in the research, we employed the Glarborg et al. [119] mechanism to investigate {\mathrm{N}\mathrm{H}}_3 oxidation at elevated pressure. Our findings, which are depicted in Fig. 20 [37], suggest that the temperature range that is most favorable for {\mathrm{N}\mathrm{H}}_3 oxidation decreases as the pressure rises. It is possible that the acceleration of ammonia-consumption reactions-such as Eqs. (24)-(27)- during combustion is the reason for this observed shift toward lower temperatures for ideal {\mathrm{N}\mathrm{H}}_3 oxidation with increasing pressure. Another possible explanation for the shift in the optimum temperature window in Fig. 20 [37] is the decrease in ignition delay time caused by high pressure, which leads to lower oxidation temperatures. This hypothesis is consistent with previous research [123,124]. Furthermore, NO concentration drops as pressure rises, suggesting that high pressure promotes \mathrm{N} \mathrm{O} reduction through the main reactions as Eqs. (28)-(30) show.

Char-N oxidation in pressurized air combustion was studied by Lin et al. [125], who discovered that the emissions of {\mathrm{N}\mathrm{O}}_x decreased significantly when the pressure increased, due to hindered {\mathrm{N}\mathrm{O}}_x diffusion from the char particles. Fig. 21 [125] shows how the oxidation of char-N is affected by pressure and temperature, with an oxygen concentration of up to 5 \mathrm{\%} in the flue gases and a remaining balance gas of {\mathrm{N}}_2. With increasing temperature, the oxidation of char-N to {\mathrm{N}\mathrm{O}}_x increased at atmospheric pressure; at a higher pressure, however, the conversion of char-N to {\mathrm{N}\mathrm{O}}_x decreased. Under POC conditions, according to Liang et al. [115], the reduction of {\mathrm{N}\mathrm{O}}_x on the char surface is facilitated by {\mathrm{C}\mathrm{O}}_2 gasification, which results in CO production and helps to reduce NO emissions.

Over time, there has been an increased need to minimize {\mathrm{N}\mathrm{O}}_x emissions from solid fuel combustion. Various conventional methods have been employed to address this problem, including fuel staging, air staging, SCR, SNCR, internal recirculation, and flameless combustion [35]. However, none of these {\mathrm{N}\mathrm{O}}_x -control technologies have been tested under high-pressure conditions.

The SNCR method is a practical and effective way to lower {\mathrm{N}\mathrm{O}}_x emissions in flue gases [126]. This method involves adding reducing reagents-such as ammonia, urea, or cyanuric acid-to the flue gas stream carrying {\mathrm{N}\mathrm{O}}_x. Within a temperature range of 1050- 1300 \hspace{0.33em} \mathrm{K}, certain reagents can reduce {\mathrm{N}\mathrm{O}}_x to {\mathrm{N}}_2 through a homogeneous reaction, as described in Refs. [27,127]. The SNCR process offers several benefits compared with catalytic processes. Multiple studies have examined the detailed chemistry of the SNCR method at atmospheric pressure. Although SNCR has several benefits over catalytic {\mathrm{N}\mathrm{O}}_x reduction methods, its main drawback is the fact that its ability to reduce {\mathrm{N}\mathrm{O}}_x is limited to within a temperature range of 1050 - 1300 \hspace{0.33em} \mathrm{K} . {\mathrm{N}\mathrm{O}}_x reduction reactions have an insignificant rate when the temperature is below 1050 \hspace{0.33em} \mathrm{K}, and {\mathrm{N}\mathrm{O}}_x formation becomes the dominant process when the temperature is above 1400 \hspace{0.33em} \mathrm{K}. To comprehend the performance of the SNCR process under high pressure, we utilized a gas-phase mechanism derived from GRI Mech 3.0, which included a subset of nitrogen chemistry from Glarborg et al. [57] and Mueller et al. [128]. The mechanisms were validated according to the experimental results by Dagaut et al. [129].

Fig. 22 [27] depicts how pressure affects the SNCR process at different temperatures. The reported results show that increasing the pressure from 0.1 to 1.5 \mathrm{M} \mathrm{P} \mathrm{a} leads to a marginal improvement in NO reduction efficiency. Furthermore, the optimal SNCR temperature window widens with increasing pressure, which increases the operational flexibility for determining the injection location of ammonia in the boiler.

The widening of the optimal temperature window in the SNCR process at a high pressure can be attributed to two potential effects: an increase in \mathrm{O} \mathrm{H} radicals and the suppression of {\mathrm{N}\mathrm{O}}_x formation reactions. These effects facilitate the transformation of {\mathrm{N}\mathrm{H}}_3 to {\mathrm{N}\mathrm{H}}_2 \left({\mathrm{N}\mathrm{H}}_3+\mathrm{O}\mathrm{H}\to {\mathrm{N}\mathrm{H}}_2+{\mathrm{H}}_2\mathrm{O}\right) and the further reaction of {\mathrm{N}\mathrm{H}}_2 with \mathrm{N} \mathrm{O} to convert it into {\mathrm{N}}_2 \left({\mathrm{N}\mathrm{H}}_2+\mathrm{N}\mathrm{O}\to {\mathrm{N}}_2+{\mathrm{H}}_2\mathrm{O}\right).

The reburning process is a technology that can be used to decrease {\mathrm{N}\mathrm{O}}_x emissions. In this process, fuel staging is used to control {\mathrm{N}\mathrm{O}}_x [130]. In the initial step, most of the fuel is burned with excess air, resulting in the formation of a sizable quantity of {\mathrm{N}\mathrm{O}}_x. In the second stage, 15 \mathrm{\%} - 25 \mathrm{\%} of the reburning fuel is added to create a fuel-rich atmosphere in which hydrocarbon radicals are produced. These hydrocarbon radicals react with {\mathrm{N}\mathrm{O}}_x and reduce the {\mathrm{N}\mathrm{O}}_x to {\mathrm{N}}_2 [119]. After that, an over-fire air system is added to combust the remaining unburned hydrocarbon radicles. This process can decrease {\mathrm{N}\mathrm{O}}_x emissions by 50 \mathrm{\%} - 70 \mathrm{\%} [131]. Previous studies have shown that the modeling of fuel-NO reburning at atmospheric pressure has generally achieved acceptable results. However, no research has investigated the reburning process at high pressure thus far.

A study conducted by \mathrm{Z} \mathrm{u} et al. [38] validated multiple mechanisms under fuel-rich conditions, ultimately selecting PG 2018 as the most appropriate mechanism for studying reburning at both ambient and elevated pressures. Using the Glarborg mechanism, a Chemkin simulation was conducted to investigate the influence of an increase in pressure from 0.1 to 1.5 \mathrm{M} \mathrm{P} \mathrm{a} on the reburning process; the results are presented in Fig. 23 [38]. The simulation revealed that, as the pressure increased, the concentration of {\mathrm{N}\mathrm{O}}_x decreased, with the most significant effect being observed at pressures up to 1 \mathrm{M} \mathrm{P} \mathrm{a}. This result could be due to the elevated pressure promoting the generation of {\mathrm{C}\mathrm{H}}_i and {\mathrm{N}\mathrm{H}}_i (where i is the number of \mathrm{H} atom), which quickly interact with NO. A reaction rate analysis identified the following highly active reactions that lead to the formation of these reactive radicals from the reburning fuels:

Croiset et al. [132] investigated the reduction of {\mathrm{N}\mathrm{O}}_x on char surfaces under high-pressure conditions. Their analysis revealed that the rate constant for {\mathrm{N}}_2 \mathrm{O} reduction was greater than that of NO reduction up to a pressure of 1 \mathrm{M} \mathrm{P} \mathrm{a}, so {\mathrm{N}}_2 \mathrm{O} was reduced more quickly than NO when interacting with char. This observation was supported by Rodriguez-Mirasol et al. [133]. Up to 0.6 MPa, pressure was found to have a significant impact on {\mathrm{N}\mathrm{O}}_x reduction, whereas only a minor impact was observed between 0.6 and 1.0 \mathrm{M} \mathrm{P} \mathrm{a}. An increase in the NO partial pressure leads to the production of either {\mathrm{N}}_2 or {\mathrm{N}}_2 \mathrm{O} through subsequent reactions such as \mathrm{N} \mathrm{O} + \left(-\mathrm{C}\mathrm{O}\right) \to 0.5 {\hspace{0.33em}\mathrm{N}}_2 + \left(-\mathrm{C}\mathrm{O}\right) and \mathrm{N} \mathrm{O} + \left(-\mathrm{C}\mathrm{N}\mathrm{O}\right) \to {\mathrm{N}}_2 \mathrm{O} + (-CO). Croiset et al. [132] observed that the formation of NO does not alter with a change in the total pressure; however, the total pressure affects NO reduction on the char. Therefore, an increase in the total pressure results in a decrease in overall NO emissions. Zhang et al. [41] conducted research on the structure of char in pressurized pyrolysis, utilizing {\mathrm{C}\mathrm{O}}_2 as the balance gas. They discovered that the gasification reaction of {\mathrm{C}\mathrm{O}}_2 greatly improved the porosity and Brunauer, Emmett, and Teller (BET) surface area of the char. This finding indicates that the reduction of {\mathrm{N}\mathrm{O}}_x through heterogeneous reactions with the char may be greatly increased at higher pressures. As shown in Fig. 24, Zhang et al. [39] investigated the reduction of NO by char in a pressurized DTF and discovered that increasing the pressure increased the NO reduction efficiency. The high-reactivity char produced in a pressurized DTF had a greater reduction performance than the char produced in a pressurized fixed-bed reactor with a low heating rate. However, more studies are needed to elaborate the influence of elevated pressure on the reduction of NO by char.

Ash aerosol formation is a concern in POC. During coal combustion, the ash contents undergo vaporization, condensation, fragmentation, agglomeration, and so forth. As a result, fine PM is emitted from the combustor, seriously threatening both human health and visibility [134]. Moreover, ash deposits on heat-exchanging surfaces may impair heat transfer or even cause dangerous accidents [135]. Here, reviews on PM formation in pressurized pyrolysis and combustion are respectively provided in Sections 4.1 and 4.2, while PM control strategies are summarized in Section 4.3.

Char structure plays a vital role in fragmentation and ash evolution during burnout [136]. It has been widely reported in previous decades that the morphology and structure of the produced char can be significantly influenced by the pyrolysis pressure, especially for bituminous coal [137]. Based on char cross-section and morphology, Wall et al. [136] developed a simplified char classification system that divides char into three groups: I, II, and III. Char particles in group I have a thin wall with high porosity \left(>70\mathrm{\%}\right), char particles in group III have low porosity , and char particles in group II have medium porosity and wall thickness. An early study by Wu et al. [137] reported that coal pyrolyzed at higher pressures produced much more char belonging to group I than coal pyrolyzed at 0.1 \mathrm{M} \mathrm{P} \mathrm{a}. Since then, researchers have attempted to investigate the interaction of coal pyrolysis factors (atmosphere, heating rate, coal type, etc.) with pressure [17,138]. A recent pyrolysis study on two ranks of coal (bituminous and lignite) was carried out in a pressurized EFR [41]. Fig. 25 [41] shows the morphology of bituminous char produced in two atmospheres \left({\mathrm{N}}_2\right. or {\mathrm{C}\mathrm{O}}_2 as the balance gas, and vice versa) at elevated pressures ( 0.1 , 0.4, 0.7, and 1.0 \mathrm{M} \mathrm{P} \mathrm{a}). The bituminous char morphology showed a significant change at increased pressure: Char samples produced at atmospheric pressure were denser; in comparison, in the char produced at elevated pressure, char swelling was obvious, the wall thickness of the produced char was decreased, and a number of pores were distributed on the particle shell. However, no significant change was found in lignite char produced at varied pressures. In addition, only minor differences in char morphology were found in an {\mathrm{N}}_2 versus {\mathrm{C}\mathrm{O}}_2 atmosphere for different ranks of coal.

Swelling is the main contributor to the pressure affecting char evolution. A number of works have explored the relevance of pyrolysis pressure and softening coal swelling [137,139,140]. Fig. 26 [137,139,140] shows trends in the swelling ratios of various types of coal with increased pressure. In general, swelling can be significantly promoted by an increase in pressure. However, there is an optimized pressure at which the swelling ratio reaches the maximum value; at pressures past that point, the influence of pressure on coal swelling is weakened.

While the factors related to coal swelling are complicated and interactive, they are generally determined by the pressure, heating rate, and coal type [141]. A great deal of studies have attempted to describe the correlation between pressure and swelling ratio using empirical models [142]. Typically, these models have been fitted by pyrolysis pressure and coal properties (proximate, ultimate analyses, etc.), but the heating condition and intrinsic swelling mechanism have not been considered. Bubble models, which can take detailed swelling mechanisms into account, have been employed to predict coal swelling. To simulate the softening coal evolution at atmospheric pressure, Oh et al. [143] were the first to propose a multi-bubble model that included detailed bubble evolution behaviors. Yu et al. [144] simplified the multi-bubble model and validated it using bituminous coal with different densities; they revealed that the modified model is applicable for coal evolution prediction in a wide range of heating rates and pressures. The bubble model was then improved by Yang et al. [145] to also estimate the specific surface area. Unlike the multi-bubble model, the single-bubble model assumes that only a single bubble is generated during coal devolatilization. Based on the work of Solomon et al. [146] and Sheng and Azevedo [147], Zhang et al. [42] modified the single-bubble model, validated it, and applied it to calculate the structure development of softening coal during pyrolysis over a wide range of pressures and heating rates.

In bubble models, the pyrolysis process can generally be divided into three stages; the pre-plastic and solidified stages can be treated as volumetric reactions, while the main swelling process occurs in the plastic stage. The division between stages is based on the viscosity of the coal melt and depends on the particle metaplast content and temperature. At elevated pressure, volatile matter-especially tar products-is significantly suppressed; as a result, the volatile-release rate is decreased, while the metaplast content is increased. Fig. 27 [42] shows the coal melt viscosity during heating based on this correlation. Due to the higher metaplast content, the particle viscosity at higher pressures is lower at 875 - 950 \hspace{0.33em} \mathrm{K}, which can contribute to coal swelling.

In coal combustion, a combusting particle yields many fragments, rather than one single particle. This phenomenon indicates that intense breaking occurs in the char consumption process [148]. It has been demonstrated that the char generated at elevated pressure undergoes swelling, as a result, the porous particles are more prone to be fragmented. Our previous work quantitatively estimated the capacity for fragmentation of char produced in a pressurized environment [40]. In that work, char samples were collected through a vertical reactor at a specific pressure, while the char combustion experiment was completed in an EFR at atmospheric pressure, in order to maintain the same heating and combustion environment for the varied char samples. Previous research has reported three categories of PM: the fine, central, and coarse modes. PM mode division is based on the sum of Si and \mathrm{A} \mathrm{l} mass fractions in the PM. Researchers demonstrated that the total mass fraction of \mathrm{S} \mathrm{i} + \mathrm{A} \mathrm{l} in fine-mode and coarse-mode PM were nearly independent of particle size, while the \mathrm{S} \mathrm{i} + \mathrm{A} \mathrm{l} mass fraction in central-mode PM exhibited a linear dependence on the particle size [149]. As shown in Fig. 28 [40], the combustion of char collected at a higher pressure generally yields more {\mathrm{P}\mathrm{M}}_{10}, while the effect of the pyrolysis atmosphere on the PM size distribution can be neglected. In particular, for char samples produced in a {\mathrm{C}\mathrm{O}}_2 atmosphere, the fine-mode, middle-mode, and coarse-mode PM derived from 0.7 \mathrm{M} \mathrm{P} \mathrm{a} char combustion respectively increased by 54.5 \mathrm{\%} , 54.7 \mathrm{\%}, and 55.9 \mathrm{\%} in comparison with those from 0.1 \mathrm{M} \mathrm{P} \mathrm{a} char.

Although a variety of advanced equipment is available for obtaining the characteristics of PM, most instruments-such as the low-pressure impactor (LPI) [150], scanning mobility particle sizer (SMPS) [151], and electric LPI (ELPI) [152]-are designed to work at atmospheric pressure. Therefore, in order to measure PM emissions from a pressurized facility, the sampling gas generally needs to be reduced to atmospheric pressure. In general, such facilities include a pressurized DTF, FB, gasifier, and pilot-scale furnace, while the studied pressure ranges from 0.1 to 3.0 \mathrm{M} \mathrm{P} \mathrm{a}, as shown in Table 3 [154⇓⇓⇓⇓-159]. The sampling method for PM emissions in a pressurized facility is under dispute. Wu et al. [153] reported a PM sampling system that used a valve for depressurization and a cyclone for collecting ash particles larger than 2 \mu \mathrm{m}, while a filter was used to collect finer ash. The PM size-distribution measurement was offline and post-processed by image tools. Other researchers have applied more sophisticated instruments to measure particle-size distribution, whether online or offline. Valves are commonly used in a pressurized system for depressurization. However, the channel passing through such valves is complex, and ash particles present in the flow-especially large particles-tend to be trapped in the flow channel. As a result, the flow channel can be partially blocked by trapped particles, which may be amplified at higher pressure. In contrast, finer particles tend to pass through the channel continuously. To reduce the PM measurement bias caused by the depressurization process, critical nozzles were designed in some studies to control the sampling gas flow rate. The inner flow path of a critical nozzle is smooth, avoiding an abrupt change in the cross-section and thereby minimizing particle losses and the frequency of channel blockages.

Limited measurement results are available regarding the PM distribution from coal combustion at elevated pressure. Wu et al. [153] investigated the aerosol emissions of bituminous coal inside a pressurized DTF and reported the influence of pressure on ash size distribution. As shown in Fig. 29 [153], four pressures-namely,0.1,0.5,1.0, and 1.5 \mathrm{M} \mathrm{P} \mathrm{a} - are compared. Ash emitted at a higher pressure tends to be much finer than that emitted at a lower pressure. Wu et al.’s study [153] confirmed that pressure influences ash formation (especially particles larger than 1 \mu \mathrm{m}) by its influence on the char characteristics after pyrolysis.

The PM emissions in oxy-coal combustion have been investigated in different facilities, ranging from the lab to pilot scales. Although the current knowledge of the effect of pressure on PM evolution is very limited, many researchers have focused on PM evolution under atmospheric-pressure oxy-fuel conditions [160⇓- 162]. An environment with a high {\mathrm{C}\mathrm{O}}_2 partial pressure may suppress the release of metal oxides in the gas phase and reduce the formation of sub-micron particles.

PM formation in oxy-combustion has been widely studied. Jia and Lighty [163] studied PM formation in both air combustion and oxy-combustion. As shown in Fig. 30 [163], an obvious decrease in the yield of ultrafine PM was found when the balance gas was changed from {\mathrm{N}}_2 to {\mathrm{C}\mathrm{O}}_2 with the same volume fraction of {\mathrm{O}}_2. Compared with an {\mathrm{N}}_2 atmosphere, the oxygen diffusion rate to the char surface in a {\mathrm{C}\mathrm{O}}_2 atmosphere is lower, which may inhibit the char oxidation rate. However, negligible differences were found in the yield of coarse-mode PM. A similar comparison was conducted by Ruan et al. [164] in an EFR: Three kinds of Zhundong lignite coal were selected to study the influence of the combustion atmosphere on {\mathrm{P}\mathrm{M}}_{10} formation. The work revealed that the {\mathrm{P}\mathrm{M}}_1 yields from oxy-combustion were much lower than those under traditional air conditions, even when the {\mathrm{O}}_2 volumetric fraction reached 30 \mathrm{\%}. Although the same gas temperature was maintained under all conditions, the temperature of the burning particles and surrounding flue gas was lower in an oxy-environment than in air combustion because of the high specific heat and intense heat radiation property of flue gas in an oxy-environment. Accordingly, the vaporization of inorganic elements is weakened in oxy-combustion. However, this difference is gradually diminished as the volumetric fraction of {\mathrm{O}}_2 increases from 30 \mathrm{\%} to 40 \mathrm{\%} because of the enhanced oxidation rate. Zhan et al. [165] investigated PM formation in a 100 \hspace{0.33em} \mathrm{k} \mathrm{W} oxy-fuel combustor; they found that, when the flame temperatures in the radiant zone were maintained, oxy-coal combustion had no effect on the deposit compositions, deposition rates, or ash aerosol properties in comparison with air combustion.

A comparative study on PM formation from air combustion and oxy-combustion was conducted in a 25 \hspace{0.33em} \mathrm{k} \mathrm{W} self-sustained furnace. Although the furnace temperature was maintained at a similar level in different combustion modes, the mass distribution of {\mathrm{P}\mathrm{M}}_{10} showed distinct differences. The researchers found that the {\mathrm{P}\mathrm{M}}_{10} and {\mathrm{P}\mathrm{M}}_1 concentrations formed in oxy-combustion were much higher than those in traditional air combustion; they attributed this finding to the higher vaporization and nucleation in oxy-combustion with a 30 \mathrm{\%} {\mathrm{O}}_2 mole fraction.

In the early stage of coal combustion, both mineral vaporization and soot generation are dominating paths for fine-particle formation [166]. An amount of ultrafine PM can be observed in the flame region. The interaction between vaporizable minerals and soot is also a concern. Soot is usually yielded when the combustion environment is sufficiently fuel-rich to allow heavy hydrocarbon condensation or polymerization [167]. However, in POC, the reaction intensity between oxidants and fuels can be significantly enhanced because of the narrowed reaction zone. To prevent overheating around the burner and to form a uniform heat flux along the length of the furnace, maintaining a low equivalence ratio and organizing a diffusion flame is recommended [14,168]. Thus, the evolution of fine-mode PM in the initial stage during POC in a full-scale furnace is of concern. In view of the risk of PM emissions, the soot formation and fine-mode PM in the early stage of combustion-as well as those in char burnout-should be considered when determining the oxidant and fuel distribution.

In the previous decade, the sub-micron PM evolution mechanism during POC was revealed using online techniques. It was acknowledged that the vaporization-nucleation-condensation mechanism of the minerals inside particles contributes to the formation of fine-mode \mathrm{P} \mathrm{M} . {\mathrm{P}\mathrm{M}}_1 formation can be influenced by many factors, including combustion pressure, temperature, coal type, facility, fuel preheating, additives, and so forth. Khatri et al. [159] demonstrated pressurized oxy-coal combustion in a 100 \mathrm{k} \mathrm{W} self-sustained combustor with an optimized sampling system and estimated the evolution of {\mathrm{P}\mathrm{M}}_1 at 1.5 \mathrm{M} \mathrm{P} \mathrm{a}, as shown in Fig. 31. The peaks of the {\mathrm{P}\mathrm{M}}_1 particle-size distribution tend to decline with an increase in residence time. The researchers attributed this phenomenon to the coupling influence of soot consumption and ultrafine PM agglomeration. Based on a lab-scale pressurized FB, Duan et al. [157] studied the effect of pressure on PM size distribution and mineral migration in a 10 \hspace{0.33em} \mathrm{k} \mathrm{W} pressurized FB, as shown in Fig. 32. It was demonstrated that, as the operating pressure increased, the concentration of ultrafine particles decreased accordingly, which was attributed to the promotion of carbon consumption at a higher pressure. Moreover, the increase in pressure may promote the vaporization and release of alkali and alkaline earth metals; as a result, the partitioning of alkali and alkaline earth metals inside ultrafine particles has been found to be higher under high pressure than atmospheric pressure. Using the same type of FB, Qiu et al. [169] studied the effect of pressure on ash particle size distribution. Their results indicated that the mass concentration of {\mathrm{P}\mathrm{M}}_1 clearly decreased at higher pressure. The researchers noted that agglomeration was an important mechanism of fine ash formation during POC. As the pressure was elevated, the mass flow rates of both the oxidant and the fuel increased linearly; accordingly, the mass concentration of the local ash increased, with more fine particles being converted into coarse-mode PM. Li et al. [158] investigated the effects of pressure on the particle size distribution of sub-micron aerosols in a pilot-scale pressurized combustor; their results showed that the submicron fraction of the total ash yield was lower at higher pressure, even though the flame temperature was carefully maintained at the same level in the different cases. Aside from the agglomeration mechanism, the researchers speculated that metal vaporization might be inhibited at elevated pressure, which could diminish the initial nucleation of sub-micron particles.

Fine particulate control during combustion has been intensively investigated in previous research. Although most of the theories are based on traditional combustion, they are applicable to POC after some modification. There is a consensus that the removal efficiency of PM in POC-especially that of sub-micron particles-is insufficient [170], due to the characteristic of the particles’ intrinsic charging capacity. Accordingly, sub-micron PM removal control is important in coal-fired power plants. In general, control strategies for the PM derived from coal combustion can be classified into two categories: combustion optimization in the reaction zone and post-combustion control technology [171]. The potential performance of these strategies when applied in POC is summarized below.

Injecting additives is an effective approach to remove fine PM in the flame zone. Kaolinite, attapulgite, and calcium hydroxide are the most frequently reported additives and have been demonstrated to be efficient in ultrafine PM removal [172]. It has been shown that injecting additives into the flame zone can suppress the conversion of vaporizable minerals into fine PM and convert them into large particles in the furnace that can be easily removed. The chemical and physical absorption of metal vapor and PM nuclei are concurrent during the capture process. Thus, less fine or ultrafine PM enters the dust collectors when additives are injected, improving the performance of dust collectors. Although no valid experiment has been conducted yet to study the performance of additives in POC, it is suspected that the above mechanism will be significantly enhanced at elevated pressure. Intrinsically, if the ratio of additives to fuel remains constant, as the pressure increases, the reaction probability between the additives and metal vapors will linearly increase; as a result, the heterogeneous reactions induced by the additives will be significantly enhanced. This mechanism was validated in our previous work in regard to {\mathrm{N}\mathrm{O}}_x removal using coal/char in POC [173].

Regarding post-combustion technologies, the electrostatic precipitator (ESP), bag filter (BF), and their combination are dominant in coal-fired power plants, due to their advantage of high efficiency [39,174]. Considering the relatively low capture efficiency of {\mathrm{P}\mathrm{M}}_{2.5} in traditional systems, novel agglomeration technologies such as magnetic agglomeration, acoustic agglomeration, and turbulent agglomeration have been developed to pre-process fine particles before they enter the ESP/BF system [175]. Compared with traditional combustion under atmospheric pressure, the furnace space is significantly compressed in POC, making it challenging to arrange complicated PM collectors in a narrow space. In addition, due to the high pressure, the risk of leakage increases, which may cause the release of fine particles directly into the environment. There is a lack of experiments and theory regarding PM capture in a pressurized system at the post-combustion stage, and further basic research is required in the future.

The mercury in coal bounds to various mineral forms, including pyrite, sphalerite, and galena. Organic mercury and mercury chloride are also found in some coal species [176,177]. The mercury migration and transformation process during coal combustion is illustrated in Fig. 33. In the process of coal combustion, a high temperature can lead to the complete decomposition of every type of mercury compound. In the flue gas flow process, when the temperature of the flue gas drops, part of the {\mathrm{H}\mathrm{g}}^0 in the flue gas is oxidized by \mathrm{H} \mathrm{C} \mathrm{l} and other components in the flue gas and is converted into divalent mercury \left({\mathrm{H}\mathrm{g}}^{2+}\right). Some of the {\mathrm{H}\mathrm{g}}^0 is adsorbed onto particles, such as fly ash, to form particulate mercury \left({\mathrm{H}\mathrm{g}}^{\mathrm{p}}\right) [178].

Wang et al. [179] conducted experiments on an FB and concluded that an oxy-combustion atmosphere promoted the oxidation of mercury. It was found that, when an air atmosphere is switched to an oxy-combustion atmosphere, the concentration of total \mathrm{H} \mathrm{g} \left({\mathrm{H}\mathrm{g}}^{\mathrm{T}}\right) in the flue gas declined slightly, the proportion of {\mathrm{H}\mathrm{g}}^{2+} increased, and the concentration of {\mathrm{H}\mathrm{g}}^0 decreased slightly [180]. Diverse types of coal also result in different proportions of the three main mercury-containing substances. Sun et al. [181] studied the principles of mercury release and migration in eight typical coal washes under an air atmosphere and oxy-combustion atmosphere using a tube furnace system. Their results indicated that the {\mathrm{H}\mathrm{g}}^{\mathrm{T}} emission concentration of lignite is higher than that of bituminous coal and anthracite in both atmospheres. In an oxy-combustion atmosphere, the {\mathrm{H}\mathrm{g}}^0 in the lignite combustion flue gas accounted for about 85 \mathrm{\%} of the {\mathrm{H}\mathrm{g}}^{\mathrm{T}}, whereas the {\mathrm{H}\mathrm{g}}^0 content of the other two substances was relatively small. The main component of bituminous coal was found to be {\mathrm{H}\mathrm{g}}^{\mathrm{p}}, accounting for about 45 \mathrm{\%}. In sub-bituminous coal, {\mathrm{H}\mathrm{g}}^0 accounted for 65 \mathrm{\%} , {\mathrm{H}\mathrm{g}}^{2+} accounted for 20 \mathrm{\%}, and the remainder of the mercury was in the form of {\mathrm{H}\mathrm{g}}^{\mathrm{p}}. Low-caloric value coal (LCVC) is a type of industrial byproduct in coal mining and coal preparation in China; it includes coal gangue, slime, and some intermediate ore. The mineral content of LCVC is higher than that of ordinary coal, and the mercury content of LCVC is also higher than that of ordinary coal [182]. A relevant study [183] showed that mercury may combine with sulfide in coal, and the content of pyrite-combined mercury in LCVC was found to be positively correlated with the sulfur content. To sum up, the content of {\mathrm{H}\mathrm{g}}^0 in flue gas in an oxy-combustion atmosphere is generally high [184], similar to the mercury content in an air atmosphere, with {\mathrm{H}\mathrm{g}}^{\mathrm{P}} , {\mathrm{H}\mathrm{g}}^{2+}, and {\mathrm{H}\mathrm{g}}^0 contents of 10 \mathrm{\%} , 37 \mathrm{\%}, and 53 \mathrm{\%}, respectively. In general, most of the mercury in coal flue gas still exists in the form of {\mathrm{H}\mathrm{g}}^0, which is both the key to and a challenge in coal mercury removal.

Due to flue gas circulation, oxy-combustion flue gas contains higher concentrations of {\mathrm{S}\mathrm{O}}_2 , {\mathrm{S}\mathrm{O}}_3 , \mathrm{N} \mathrm{O} , \mathrm{H} \mathrm{C} \mathrm{l} , {\mathrm{H}}_2 \mathrm{O}, and fly ash than air-combustion flue gas. These flue gas components react with mercury in a series of homogeneous and heterogeneous phases and increase the proportion of {\mathrm{H}\mathrm{g}}^{2+} and {\mathrm{H}\mathrm{g}}^{\mathrm{p}} in {\mathrm{H}\mathrm{g}}^{\mathrm{T}} [4,185⇓⇓⇓⇓-190], which is objectively conducive to the removal of mercury via dusters and WFGD devices. However, because of the high {\mathrm{H}\mathrm{g}}^0 content in oxy-combustion flue gas, the oxy-combustion system presents more serious mercury-emission problems. A large amount of {\mathrm{H}\mathrm{g}}^0 can react with aluminum {\mathrm{C}\mathrm{O}}_2 compression and purification equipment in mercury amalgamation; this results in metal catalysis and corrosion, thus adversely affecting the safe and stable operation of an oxy-combustion system. Therefore, it is necessary to develop a mercury-removal technology that is suitable for use under an oxy-combustion atmosphere in order to ensure the secure, reliable, and environmentally friendly functioning of oxy-combustion systems.

Mercury-removal technology can be applied in three different stages of coal combustion: pre-combustion, during combustion, and post-combustion. Post-combustion mercury removal is the most extensively employed for mercury removal.

A well-designed multi-pollutant purification system can remove a large part of the mercury in flue gas. The {\mathrm{H}\mathrm{g}}^{2+} is absorbed by desulfurized slurry when passing through a WFGD unit [157]. The other part of the mercury is captured by the fly ash in the flue gas flow process and becomes {\mathrm{H}\mathrm{g}}^{\mathrm{p}} combined with PM. ESP and fabric filters (FFs) can simultaneously capture {\mathrm{H}\mathrm{g}}^{\mathrm{p}} when removing particles such as fly ash, thereby realizing collaborative Hg removal [191]. The use of existing pollutant-control equipment for collaborative mercury removal is the current research focus, as the existing equipment has a low operating cost and its use makes it unnecessary to add new equipment.

Fig. 34 illustrates the collaborative removal process for mercury in flue gas under an oxygen-enriched atmosphere via a conventional flue gas pollutant-purification device. After the flue gas flows through each heat-exchange surface, the temperature gradually drops, and part of the mercury is adsorbed onto the fly ash to generate {\mathrm{H}\mathrm{g}}^{\mathrm{p}}, thereby creating conditions for the duster to capture mercury from the flue gas. Moreover, the proportion of {\mathrm{H}\mathrm{g}}^{\mathrm{p}} in the flue gas under an oxygen-enriched atmosphere is higher than that in air-atmosphere flue gas, promoting collaborative mercury removal by ESP or FF, as well as other dust removal devices [180,192]. Mitsui et al. [192] studied the migration and transformation of mercury and {\mathrm{S}\mathrm{O}}_3 in the flue gas of a 1.5 {\mathrm{M}\mathrm{W}}_{\text{ th }} oxy-combustion experimental system. Their data showed that, when the flue gas flowed through the ESP, the mercury concentration in the flue gas decreased by about 87 \mathrm{\%}, with an obviously synergistic mercury-removal effect. The type of coal being combusted also plays a role in the synergistic mercury-removal efficiency of ESP in an oxy-combustion system. As the content of sulfur oxides in flue gas is directly associated with the sulfur content of the coal, burning high-sulfur coal increases the {\mathrm{S}\mathrm{O}}_x content in the flue gas, further inhibiting the adsorption of mercury on fly ash particles and causing the dusters’ collaborative mercury-removal performance to deteriorate.

However, in practical application, due to changes in operating conditions and the poor effect of {\mathrm{H}\mathrm{g}}^0 removal, it is impossible to achieve continuous and stable mercury removal. Thus, it is still necessary to develop a special mercury-removal technology that is economically and technically feasible and suitable for oxy-combustion flue gas.

As gaseous {\mathrm{H}\mathrm{g}}^{2+} is easily soluble in water, when it passes through a WFGD system with flue gas flow, it will be absorbed by the desulfurization slurry, along with {\mathrm{S}\mathrm{O}}_2 and other acidic gases, and enter into the desulfurization wastewater and gypsum. Using a simulated desulfurization slurry to study the migration and transformation of mercury when flowing through a WFGD device in a high-concentration {\mathrm{C}\mathrm{O}}_2 atmosphere, \mathrm{W} \mathrm{u} et al. [193] found that {\mathrm{C}\mathrm{a}\mathrm{S}\mathrm{O}}_3 could remove mercury more efficiently than {\mathrm{C}\mathrm{a}\mathrm{S}\mathrm{O}}_4. When there are many anions, such as {\mathrm{N}\mathrm{O}}_3 {}^{-} and {\mathrm{S}\mathrm{O}}_3 {}^{2-}, in the desulfurization slurry, the {\mathrm{H}\mathrm{g}}^{2+} in the desulfurization slurry is more likely to be reduced to gaseous {\mathrm{H}\mathrm{g}}^0 and released, resulting in an increase in the concentration of mercury in the final discharged flue gas. The presence of {\mathrm{C}\mathrm{l}}^{-} in desulfurization slurry is beneficial to the stable existence of mercury in the slurry, even when there are other anions present. Ochoa-González et al. [194] studied the influence of flue gas components, including {\mathrm{C}\mathrm{O}}_2 and {\mathrm{O}}_2, in air and oxy-combustion flue gas on the form of mercury in a WFGD device. The experimental results indicated that a relative high concentration of {\mathrm{C}\mathrm{O}}_2 was soluble in the desulfurization slurry, further decreasing the \mathrm{p} \mathrm{H} value and oxidation-reduction potential of the desulfurization slurry, and thereby weakening the reduction release of {\mathrm{H}\mathrm{g}}^{2+} to {\mathrm{H}\mathrm{g}}^0. In addition, the {\mathrm{O}}_2 in flue gas was found to be instrumental in the stability of mercury in the desulfurization slurry. Therefore, it is necessary to strictly control the \mathrm{p} \mathrm{H} value and oxidation-reduction potential of the desulfurization slurry so as to ensure the efficient and synergistic removal of mercury in the flue gas with an oxygen-enriched atmosphere via a WFGD device.

Adsorbents are used to remove mercury from oxy-combustion flue gas in an approach similar to that used in traditional air combustion: An adsorbent injection system for the removal of mercury is positioned in front of the dust collector, the adsorbent is injected into the flue gas to collect the mercury, and then the adsorbent is removed with fly ash. Thus far, adsorbents such as activated carbon [195,196], modified carbon-based adsorbent [197,198], modified fly ash [199,200], metal oxide [201,202], and mineral adsorbent [203] have been widely studied. In POC, the atmosphere and pressure are different from those in traditional air combustion; this may affect the performance of mercury-removal equipment, requiring the adsorbent to be optimized.

Activated carbon is the adsorbent that is most widely used to remove mercury from the flue gas produced during oxy-combustion. For practical purposes, halogen elements such as \mathrm{C} \mathrm{l} and \mathrm{B} \mathrm{r}, as well as elemental \mathrm{S}, are typically added to activated carbon. Modified activated carbon adsorbents have a better capacity for adsorbing mercury, since the components can create more active sites with a larger specific surface area on the surface of activated carbon. Lopez-Anton et al. [204] found the Hg-removal capacity of an activated carbon adsorbent to be up to 160 \mu \mathrm{g} \cdot {\mathrm{g}}^{-1} ; this was attributed to the sulfur-containing group on the surface of the activated carbon acting as a mercury-removal active site, greatly enhancing the \mathrm{H} \mathrm{g} -removal capacity of the activated carbon adsorbent. Zhuang et al. [205] investigated the efficiency of activated carbon in removing mercury from oxy-combustion flue gas. They revealed that the \mathrm{H} \mathrm{g} penetration time to the activated carbon adsorbent in the oxy-combustion flue gas was comparable to that in air-combustion flue gas, demonstrating that activated carbon can be employed for \mathrm{H} \mathrm{g} removal in an oxy-combustion atmosphere.

However, the comparatively high sulfur oxides and water vapor levels in oxy-combustion flue gas offer a challenge to the Hg-removal performance of activated carbon. The principal flue gas components that affect \mathrm{H} \mathrm{g} removal by activated carbon include high concentrations of sulfur oxides, where {\mathrm{S}\mathrm{O}}_3 has a particularly noticeable inhibitory effect on \mathrm{H} \mathrm{g} removal by activated carbon. It was discovered that, in sulfur-containing flue gas, the \mathrm{S} content on the surface of the activated carbon increases and exists as {\mathrm{S}\mathrm{O}}_3, indicating that a tiny quantity of {\mathrm{S}\mathrm{O}}_2 is adsorbed and oxidized on the activated carbon [206]. Previous studies have demonstrated that {\mathrm{S}\mathrm{O}}_3 adsorbs more strongly onto activated carbon than mercury does [207]. Thus, {\mathrm{S}\mathrm{O}}_3 blocks the pores on the surface of activated carbon and covers the active \mathrm{H} \mathrm{g} -removal sites, thereby significantly affecting the mercury-removal process. In addition, a temperature-programmed mercury desorption experiment in the same study found that new mercury compounds such as \mathrm{H} \mathrm{g} \mathrm{S} appeared on the surface of the activated carbon. These products remained on the surface of the activated carbon and occupied the original active sites of mercury removal, resulting in a continuous drop in the \mathrm{H} \mathrm{g} -adsorption capacity of the activated carbon.

A higher concentration of {\mathrm{H}}_2 \mathrm{O} in the oxy-combustion flue gas can also affect the \mathrm{H} \mathrm{g} adsorption by activated carbon. Lopez-Anton et al. [208] found that, after adding 6 \mathrm{\%} {\mathrm{H}}_2 \mathrm{O}, the \mathrm{H} \mathrm{g} removal performance of sulfur-loaded activated carbon decreased by about 15 \mathrm{\%}, compared with that of dry oxy-combustion flue gas, indicating that the water vapor decreased the number of reaction sites for adsorbing mercury on the surface of the activated carbon. In actual oxy-combustion flue gas, the maximum concentration of water vapor can be as high as 30 \mathrm{\%} ; thus, water vapor will have a more significant inhibitory effect on the \mathrm{H} \mathrm{g} removal of activated carbon [192]. Studies have also found that, in an oxy-combustion atmosphere, {\mathrm{S}\mathrm{O}}_2 has a stronger ability to inhibit \mathrm{H} \mathrm{g} adsorption onto activated carbon than {\mathrm{H}}_2 \mathrm{O} at the same concentration. However, when {\mathrm{S}\mathrm{O}}_2 and {\mathrm{H}}_2 \mathrm{O} are both present, the \mathrm{H} \mathrm{g} -removal efficiency may be improved due to the generation of the highly oxidizing {\mathrm{H}}_2 {\mathrm{S}\mathrm{O}}_4, which can oxidize mercury to {\mathrm{H}\mathrm{g}\mathrm{S}\mathrm{O}}_4[208].

The use of efficient mercury oxidation catalysts is another method for reducing high-concentration mercury emissions in oxy-combustion flue gas. In oxy-combustion, a high-efficiency SCR catalyst can be used to increase the concentration of {\mathrm{H}\mathrm{g}}^{2+} and {\mathrm{H}\mathrm{g}}^{\mathrm{p}} in the flue gas to 91 \mathrm{\%} - 96 \mathrm{\%} before the entrance of the dust collector, ensuring the efficient collaborative removal of mercury by a WFGD device [192]. Wang et al. [209] found that a high concentration of {\mathrm{C}\mathrm{O}}_2 can promote the oxidation of {\mathrm{H}\mathrm{g}}^0, although it is not conducive to oxidizing NO.

The catalytic oxidation of mercury can be further improved by doping the catalyst with other metal elements. Fernández-Miranda et al. [210] performed experiments using a \mathrm{V} / \mathrm{W} / {\mathrm{T}\mathrm{i}\mathrm{O}}_2 catalyst and a Mn-doped catalyst in an air combustion and oxy-combustion atmosphere, respectively, and found that the {\mathrm{H}\mathrm{g}}^0 oxidation performance of the doped catalyst in both the air and oxy-combustion atmosphere was improved by about 80 \mathrm{\%}. Wang et al. [202] carried out {\mathrm{C}\mathrm{e}\mathrm{O}}_2 -doping modification on a \mathrm{V} / \mathrm{W} / {\mathrm{T}\mathrm{i}\mathrm{O}}_2 catalyst and performed a mercury catalytic oxidation experiment in oxy-combustion flue gas. Their results demonstrated that the catalyst maintained stable performance during the experimental period of up to 50 \hspace{0.33em} \mathrm{h} and finally achieved a 75 \mathrm{\%} {\mathrm{H}\mathrm{g}}^0 -removal efficiency and 92.5 \mathrm{\%} NO-removal efficiency, which were higher than those of the ordinary \mathrm{V} / \mathrm{W} / {\mathrm{T}\mathrm{i}\mathrm{O}}_2 catalyst. This was attributed to the Ce doping modification significantly enhancing the oxidation and reduction ability of the catalyst and more oxygen-containing groups being adsorbed onto the surface of the catalyst to react with mercury. These oxygen-containing groups also promoted the formation of {\mathrm{N}\mathrm{O}}_2 and the homogeneous oxidation of {\mathrm{H}\mathrm{g}}^0, thus further promoting the formation of {\mathrm{H}\mathrm{g}}^{2+}.