PCNs are planar molecules with the molecular formula C₁₀H_{8- n}Cl _n; where nranges from 1 to 8 (see Figure 1). Their physicochemical properties are similar to those of polychlorinated biphenyls (PCBs) and polychlorinated dibenzo-p-dioxins and polychlorinated dibenzofurans (PCDD/Fs), exhibiting good stability, insulating properties, low flammability, and inertness. With the exception of monochloronaphthalene, which is liquid at room temperature, pure PCNs are typically colorless crystalline compounds^[16-17]. However, most industrially produced PCNs are not pure substances but mixtures composed of various congeners. The physical state of commercial PCN products ranges from low-viscosity liquids to hard waxes and high-melting-point solids, with melting points ranging from -40 to 180 ℃^[18].

The physicochemical properties of PCN congeners vary significantly depending on the degree of chlorine substitution. Water solubility, vapor pressure, and the Henry's law constant (K _H) decrease progressively from monochloronaphthalene to octachloronaphthalene, while the octanol–water partition coefficient (K _ow), melting point, and boiling point increase progressively from monochloronaphthalene to octachloronaphthalene^[19]. PCNs are primarily soluble in organic solvents such as benzene, dichloromethane, ether, hexane, isooctane, petroleum ether, and toluene^[16]. Trichloronaphthalene through octachloronaphthalene are highly lipophilic, with high K _owvalues generally greater than 5. The octanol–air partition coefficient (K _oa) shows a clear increasing trend with increasing levels of chlorination. From dichloronaphthalene to hexachloronaphthalene, the K _oavalues increase by more than three orders of magnitude^[20]. PCNs exhibit a strong absorption peak at 220–275 nm and a weaker absorption peak at 275–345 nm. As the degree of chlorination increases, the wavelength of maximum absorption shifts to longer wavelengths^[16]. Some physicochemical parameters of PCNs are shown in Table 1.

PCNs exhibit toxicity characteristics similar to those of dioxins, with their toxicity being more closely associated with their planar structure and β-chlorine substitution on the naphthalene ring (Figure 1) ^[21]. From monochloronaphthalene to octachloronaphthalene, all congeners have persistent adverse effects on living organisms^[22]. Some PCN congeners have structures similar to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and exhibit comparable toxic effects, including embryotoxicity, hepatotoxicity, immunotoxicity, skin damage, and carcinogenicity^[21,23]. Hepatotoxicity primarily manifests as hepatomegaly, fatty liver, and steatosis; in severe cases, it can lead to acute hepatic atrophy and cirrhosis^[18,24]. In terms of neurotoxicity, PCNs affect γ-aminobutyric acid (GABA) metabolism, leading to decreased levels of GABA and glutamate, which in turn impairs motivational processes (manifesting as appetite-suppressing effects associated with difficulty swallowing—aphagia—and lack of thirst—adipsia), reduced motor activity (hypokinesia), long-term memory impairment, and a marked attenuation of stress responses^[23]. Prenatal and developmental toxicity includes fetotoxicity, embryotoxicity, and teratogenicity, potentially resulting in skeletal dysplasia, renal abnormalities, and neurological developmental disorders^[23]. In terms of endocrine disruption, PCNs exhibit estrogenic/anti-estrogenic and androgenic/anti-androgenic effects, potentially interfering with sex hormone function and causing irregular estrous cycles and elevated levels of luteinizing hormone and follicle-stimulating hormone^[25].

In addition, similar to the mechanism of action of TCDD, PCNs can cross the cell membrane and bind to the aryl hydrocarbon receptor (AhR)^[17]. After ligand binding, the AhR-ligand complex translocates to the nucleus and forms a dimer by binding to the AhR nuclear transcriptional activator^[26]. The resulting dimer has high affinity and can bind to xenobiotic response elements (XREs) in specific target genes, thereby upregulating the transcription of certain genes (such as certain genes in the cytochrome P-450 (CYP) family). In particular, CYP genes (such as CYP1A1, CYP1A2, and others) typically play an important role in metabolizing exogenous compounds. In this process, mutagenic intermediates may form, which could lead to the initial formation of tumors^[23].

As a typical class of POPs, PCNs exhibit significant persistence, bioaccumulation, and long-range transport in the environment. The persistence of PCNs in the atmosphere is influenced by the degree and position of chlorine substitution. With the exception of monochloronaphthalene, all PCNs have an atmospheric half-life exceeding 2 days^[19]. Among the congeners, those in which all α-positions are substituted with chlorine are more persistent, whereas congeners containing the -C(β)Cl-C(α)H-C-C(α)Cl- fragment exhibit lower persistence, as the electron density on the α hydrogen and carbon atoms in this fragment is relatively low^[27]. Due to the absence of hydrolyzable functional groups in their chemical structure, PCNs are resistant to hydrolysis but can undergo photodegradation^[17]. The photodegradation of PCNs is not limited to any specific environmental conditions and can be observed even at low altitudes^[28]. In addition, the biodegradability of PCNs is also relatively limited. Studies have shown that certain low-chlorinated congeners (such as Mono-CN1 and Mono-CN2) can be degraded under the action of specific microorganisms, such as white-rot fungi^[29]. In natural environments, the degradation of PCNs is extremely slow; for example, analysis of soil and sediment samples indicates that the half-life of trichloro- to heptachloronaphthalenes exceeds one year^[19].

The lipophilicity of PCNs enables them to bioaccumulate in living organisms and undergo biomagnification through the food chain. As the degree of chlorination increases, so does the extent of bioaccumulation of PCNs. For example, in fish, the bioconcentration factor for monochloronaphthalene is 191, while that for tetrachloronaphthalene can reach as high as 34,000, demonstrating an exceptionally strong bioaccumulation capacity. However, octachloronaphthalene exhibits a relatively lower degree of bioaccumulation due to its unique chemical structure^[18]. During the bioaccumulation process, different congeners of PCNs display significant differences. Low-chlorinated congeners predominate in organisms at lower trophic levels, such as zooplankton and small fish, whereas the proportion of highly chlorinated congeners increases in organisms at higher trophic levels, such as large fish, marine mammals, and birds. This distribution pattern is closely related to the metabolic capabilities of PCNs. Some PCN congeners, due to the presence of ortho carbon atoms that are not substituted by chlorine, are more readily metabolized or excreted by organisms. In contrast, isomers with unchlorinated ortho carbon atoms (such as Hexa-CN66/67 and Penta-CN54) tend to accumulate more readily in organisms, particularly in those at higher trophic levels^[11].

The long-range environmental transport capacity of PCNs is a key reason for their global distribution^[30]. Less chlorinated congeners, due to their higher volatility, can undergo long-range atmospheric transport via global atmospheric circulation, resulting in their widespread distribution worldwide^[20]. In contrast, more highly chlorinated congeners are transported over long distances primarily through oceanic pathways, with relatively limited atmospheric transport^[20]. Monitoring data further confirm the widespread occurrence of PCNs in polar environments. Harner et al.^[31]reported the presence of PCNs in Arctic air and indicated that the predominant PCN congeners in the atmosphere are trichloronaphthalene and pentachloronaphthalene. Dong et al.^[32]analyzed PCNs in benthic marine sediments and various organisms collected near Chinese polar research stations on Svalbard in the Arctic and the South Shetland Islands in the Antarctic. The results showed that in sediments from the Arctic and Antarctic regions, the total mass fraction of ∑PCNs (75 congeners), expressed as dry weight (dw), was 274–837 pg/g (average 485 pg/g) and 72–243 pg/g (average 142 pg/g), respectively; in benthic marine biota, the total mass fraction of ∑PCNs (75 congeners), expressed as dry weight (dw), was 37–388 pg/g and 53–852 pg/g. Dong et al.^[13]analyzed the concentrations and distribution of PCNs in Arctic soils and plants on the Svalbard archipelago, finding that in soils, the mass fraction of ∑PCNs (75 congeners), expressed as dry weight (dw), ranged from 5.3 to 2550 pg/g, while in plants, the mass fraction of ∑PCNs (75 congeners) ranged from 77 to 870 pg/g.

PCNs were first synthesized in 1833 and patented in the United States in 1909^[50]. In the early 20th century, the first patented application of PCNs was as a flame retardant and dielectric fluid in capacitors^[43]. During World War I, PCNs were first used as an important protective coating material, particularly in Germany. After World War II, PCN production began to decline, gradually being replaced by plastics and PCBs used for insulation purposes. Historically, the main producing countries of PCNs have included the United States, Germany, the United Kingdom, France, Italy, Japan, Poland, and Russia (Table 2).

Different studies exhibit significant discrepancies in their estimates of global PCN production, primarily due to variations in estimation methods, with some studies failing to describe their specific methods. Falandysz et al.^[21]estimated total PCN production at 100 t by multiplying the global total production of PCB products (approximately 1.5 million tons) by the median concentration of PCNs in PCB products (0.0067%); AMAP^[51]did not specify its estimation method, stating that PCN production ranged from 20,000 to 40,000 tons; Jakobsson et al.^[16]indicated that PCN production in the 1920s was approximately 9,000 tons, without citing specific sources; Bogdal et al.^[52]proposed the assumption that "PCN production does not exceed 10% of PCB production," estimating PCN production at 130,000 tons; UNEP^[53]indicated that PCN production ranged from 15,000 to 40,000 tons.

As the primary producer of PCNs, the United States began PCN production in 1912, and the last U.S. PCB manufacturer, Chemisphere, ceased production in 1980^[18]. However, according to a 1975 report by the U.S. Environmental Protection Agency (EPA), the only U.S. PCN manufacturer was Koppers Company, which produced PCNs under the trademark Halowaxes at a plant in Bridgeville, Pennsylvania^[55]. Koppers Company ceased production in 1977^[18]. Due to the lack of early chemical production statistics, information on early PCN production is scarce, and some descriptions are inconsistent. According to Klimczak et al.^[1], from 1912 to 1937, annual PCN production in the United States was approximately 1 million pounds (460 t). In the 1920s (1920–1930), annual production surged to 20 million pounds (9,100 t), representing a substantial increase during this period. Furthermore, between 1910 and 1960, total PCN production in the United States ranged from 50,000 to 150,000 tons. According to the EPA’s 1975 report^[55], PCN production in 1956 was 7 million pounds (3,200 t), declining to 5 million pounds (2,300 t) by 1972, and further dropping to 700,000 pounds (320 t) by 1978.

Compared with global production, the amount of PCNs produced in Poland is relatively small. Production activities are primarily carried out by the Zakłady Azotowe-Mościce Group, located in the Mościce-Tarnów industrial area in southeastern Poland. The group uses gaseous chlorine and FeCl₃as a catalyst to chlorinate molten naphthalene at 130 ℃ for several hours until the desired mass density is achieved. The resulting mixture is then distilled under reduced pressure, yielding Mono-CN1 in liquid form, while the solid product is Woskol. The annual production of Mono-CN1 is 15 t, and the annual production of highly chlorinated (three to eight-chloronaphthalene) Woskol is 9 t. Between 1936 and 1939, the total PCN production in Mościce amounted to 47 t of Mono-CN1 and 36 t of Woskol. In 1947–1948, the Mościce chemical plant was partially rebuilt following the destruction sustained during World War II. Production of Mono-CN1 and Woskol was resumed and used in a product called Xylamit (a wood preservative), which was produced from 1950 to 1987. It is estimated that between 1951 and 1987, the total production of Woskol was 369 t, and the total production of Mono-CN1 was 602 t^[1].

From 1940 to 1976, Japan produced approximately 4,000 tons of PCNs and banned their production and use in 1979^[56]. Germany did not cease production until 1983; according to the United Nations Environment Programme (UNEP), Bayer produced 100–200 tons of PCNs annually from 1980 to 1983 before ceasing production in 1983^[57]. The UK stopped production in the mid-1960s, but small quantities of PCNs were still produced in the 1970s^[43]. Since 1919, ICI (Runcorn) and its predecessors have been producing Seekay waxes. In Russia, in addition to the production of monochloronaphthalene, trichloronaphthalene, and highly chlorinated naphthalenes in Usolie-Sibirskoe, the Orgsteklo and Orgsintez plants also produced PCNs^[1].

Historical emission studies of PCNs have primarily employed analytical methods such as sediment cores, tree rings, and integrated environmental media, making significant progress in elucidating emission patterns and sources. Researchers collect sediment core samples from different depths and use radiometric isotopes (e.g., ²¹⁰Pb, ¹³⁷Cs) to determine the age of sediments, thereby establishing a temporal relationship between PCN concentrations and time. For example, Gevao et al.^[58]found that the PCN flux in Esthwaite Water Lake in the UK was stable before the 1940s, peaked in the late 1950s to early 1960s, and then gradually declined. Horii et al.^[59]also reported that in Lake Kitaura, PCN concentrations peaked in the early 1980s and subsequently decreased. Yamashita et al.^[60]studied PCN emissions in Tokyo Bay and found that emissions increased from the 1960s, peaked in 1980, and then declined; the sources were associated with combustion processes and highly chlorinated PCN products. Studies in Jiaozhou Bay indicate that PCN emissions began to increase in the 1950s, peaked in the early 1990s, and then declined, closely linked to wood and coal combustion, Halowax inputs, and the incineration of municipal solid waste^[61-62]. Li et al.^[63]studied the Great Lakes in North America and found that PCN concentrations first increased and then decreased in the 20th century, with peak times varying among different lake regions. Lee et al.^[64]conducted research in Ulsan Bay, South Korea, and found that PCN emissions were closely related to local industrialization: concentrations rose and then fell between 1930 and 1980, emissions increased from the 1980s to the early 21st century, and decreased from 2006 to 2015 due to regulatory measures.

Research has found that lipophilic POPs in tree rings do not redistribute within the stem, making them historical tracers. Odabasi et al.^[65]Using tree-ring analysis, they discovered that PCN emissions from the Aliaga industrial zone in Turkey increased with industrialization starting in the 1970s. Although their production and use ceased in the mid-1980s, environmental concentrations continued to rise due to ongoing industrial emissions.

The historical emission characteristics of PCNs vary significantly across different regions. In terms of emission trends, most regions exhibit a pattern of initial increase followed by decline, but the timing of the peak and the subsequent rate of decline differ among regions. Among the influencing factors, the pace of industrialization is a key driver of changes in PCN emissions. Current research has provided some understanding of the emission patterns and pollution status of PCNs in specific local areas; however, systematic studies on global historical emissions remain relatively scarce.

In recent years, with the growing global emphasis on POPs control, research on PCN emission factors has increasingly become a focal point in environmental science. As a key parameter for quantifying pollutant emissions, emission factors provide crucial support for compiling accurate emission inventories. Major sources of PCN emissions include industrial production, metal smelting, and waste incineration. PCN emission factors exhibit significant variations across different sources, pollution control measures, and countries and regions, with numerical differences spanning up to nine orders of magnitude (Figure 4). Table 3summarizes currently published PCN emission factors. As shown in Table 3, in the steel industry, the PCN toxic equivalent quantity (TEQ) emission factor for electric arc furnaces (21.6–30.1 ng/t) is significantly higher than that for basic oxygen furnaces (0.3–1.5 ng/t). In nonferrous metal production, emission factors for secondary metal production are generally higher than those for primary metal production; for example, the TEQ emission factor for secondary copper production ranges from 35.1 to 264 ng/t, far exceeding the 0.2–13 ng/t range for primary copper production, primarily due to differences in raw materials^[68]. In contrast, PCN emission factors from household heating and biomass combustion are relatively low; for instance, the TEQ emission factors for households in the UK using hardwood and domestic coal for heating are 9 ng/t and 2 ng/t, respectively. The effectiveness of pollution control measures significantly influences PCN emission levels. For example, waste incineration plants equipped with electrostatic precipitators (ESP), wet scrubbers (WS), and selective catalytic reduction (SCR) technologies exhibit significantly lower PCN emission levels—(1.88 ± 0.472) ng/m³ under standard conditions—compared to plants equipped with cyclones (CY), semi-dry absorbers (SDA), activated carbon injection (ACI), and baghouse filters (BH), which have emission levels of (23.2 ± 0.983) ng/m³ under standard conditions^[69]. This difference is mainly attributable to SCR’s highly efficient catalytic reduction of highly chlorinated PCNs and WS’s partial ability to capture low-chlorinated PCNs. Although ESP may contribute to PCN formation, the high operating temperatures of SCR effectively decompose PCNs. In contrast, while the CY + SDA + ACI + BH combination removes PCNs through adsorption, its overall removal efficiency is lower and its emission factor is higher due to lower operating temperatures, memory effects, and the relatively high chlorine content in the waste.

In addition to the total quantity and TEQ, the congener distribution of PCNs is crucial for understanding emission characteristics, assessing toxicity risks, conducting source apportionment, and simulating environmental fate^[17]. In terms of mass emission factors, PCNs are predominantly composed of low-chlorinated CNs (Mono-CN, Di-CN, Tri-CN) (Table 3), such as CN1, CN2, and CN24/14, which have been detected at relatively high concentrations in emission factors from various sources (e.g., certain waste incinerators, magnesium smelting, and residential heating). The congener profiles of PCNs emitted from the same type of source can vary significantly across different processes, pollution control technologies, and even different countries/regions. For example, waste incineration in China using ESP+WS+SCR primarily emits CN1/CN2, whereas the CY+SDA+ACI+BH configuration mainly emits CN1^[69]; in Japan, BF+AC (activated carbon) primarily emits CN33/37^[56]; and in Germany, SDS (semi-dry scrubber)+AC+BF primarily emits CN52/60^[70]. Different stages of steel production (BOF, EAF) and different control devices (BH, WS, CY) also emit different predominant congeners (e.g., CN3, CN14/26, CN25/13, CN11/8, CN23)^[71-72]. In terms of TEQ emission factors, high-chlorinated PCNs dominate, with hexachloronaphthalene 66/67 (Hexa-CN66/67) and heptachloronaphthalene 73 (Hepta-CN73) being the primary contributors to TEQ from multiple emission sources, including waste incineration, steel production, iron ore sintering, and magnesium smelting (Table 3). These isomers have a planar structure with chlorine substitution at the β-position (consistent with the structural characteristics of dioxin-like compounds) and possess high toxic equivalency factors (TEFs)^[21]; therefore, even when their concentrations are not the highest, they often contribute the most to the total TEQ.

Although these data provide an important reference for understanding emission levels across different industries and processes, several limitations remain. In the UNEP “Guidance on Developing Inventories of PCNs” (hereinafter referred to as the “Guidance”), the emission factor for PCNs from PCB industrial mixtures is given as 39,000–1,300,000 mg/t^[53]. The emission factors are all equated to the PCN concentrations in PCB products, implying that the corresponding activity level should be the quantity of PCB products^[73]. Similarly, the emission factor of 1,300,000 mg/t mentioned in the study by Huang et al.^[74]is based on the concentration of PCNs in the insulating oil of capacitors stored with China’s main PCB product “PCB3” (also known as PCB No. 1), rather than directly measured emission concentrations. In addition, the emission factor for PCNs during carbon tetrachloride production is also equated to the product concentration^[47]. However, in the calculation process, the activity level is typically defined as the volume of product produced rather than the volume of emissions^[47,73]. This approach overlooks the emission constraints imposed by closed systems (such as insulating oil in capacitors), meaning that the actual emission levels of PCNs may be much lower than their concentrations in these systems. Furthermore, it fails to account for the complexity of PCN emissions throughout the product lifecycle—including production, use, storage, and disposal—such as the significant impact of waste product disposal methods (e.g., incineration or landfill) on the actual PCN emissions.

In addition, the emission factors for fly ash and bottom ash provided in the UNEP Guidelines^[53]use the concentration of PCNs in the ash as the emission factor, with the corresponding activity level being the emission volume of fly ash or bottom ash. However, this method has certain limitations in practical applications. The concentration of PCNs in the ash, influenced by combustion conditions, raw material composition, and pollution control technologies, may lead to deviations in emission volume estimates. More precise emission volume calculations require integration with plant operational data. For example, in cement plants, the calculation method is shown in Equation (1).

in the formula: Erepresents the PCNs emissions; Crepresents the PCNs concentration in fly ash; Mrepresents the fly ash emissions. According to the "Manual of Emission Source Statistical Survey, Emission Calculation Methods and Coefficients" issued by China's Ministry of Ecology and Environment, the calculation of particulate matter emissions during cement production must consider the following key parameters: pollution generation coefficient (the amount of pollutants generated per unit of product or raw material), product output or raw material usage (activity-level baseline data), the average removal efficiency of pollutant control technologies, and the actual operating rate of control facilities^[75].The calculation method is shown in Equation (2).

in the formula: G_prodirepresents the average particulate matter generation rate in process section i; P _prodis the pollution generation coefficient corresponding to a specific pollutant in the process section; M_iis the total output/total raw material input of process section i; R_removaliis the removal amount of a specific pollutant in process section i; η_Tis the average removal efficiency of the end-of-pipe treatment technology used for a specific pollutant in process section i; κ_Tis the actual operating rate of the end-of-pipe treatment facilities used for a specific pollutant in process section i. This multi-parameter calculation can more accurately reflect the actual emission characteristics, but it places higher demands on the comprehensiveness and accuracy of data acquisition.

Research on unintentional emissions of PCNs primarily covers waste incineration, non-ferrous metal smelting, coking, and other industrial processes. Early studies, lacking emission factors, often relied on estimates based on emission data for PCDD/Fs and PCBs. Falandysz et al.^[21]estimated that the total production of PCNs from chlorine-related thermal processes and other processes in the 20th century could be roughly estimated at 1–10 t. This conclusion is based on the global annual release of 10–100 kg of PCDD/Fs into the environment from chlorine-related thermal processes and other processes, and on the observation that the global production of PCNs from combustion and other chlorine-related processes does not appear to exceed the amount of PCDD/Fs generated. Yamashita et al.^[73]estimated PCN emissions resulting from PCB use at 169 t, based on global PCB production and the average concentration of PCNs in PCBs. Denier van der Gon et al.^[104]considered three potential sources of PCNs—waste incineration, landfilling, and PCB production—and estimated PCN emissions based on five categories: public thermal power, residential, commercial, and other combustion, industry, solvent and product use, and waste treatment. The results indicate that in 2000, the annual PCN emissions in the UNECE European region were 1 t. Of this total, waste treatment accounted for 74%, industry for 11%, and residential, commercial, and other combustion for 10%. The emission factor for industrial processes was derived by assuming an average PCN content of 0.05% ± 0.04% in PCBs, while emission factors for other sources were based on a rough estimate of an average PCN to PCDD/Fs ratio of 50:1.

In recent years, research has primarily used the emission factor method and measured data to conduct relatively systematic estimates of unintentional emissions of PCNs. Zhang et al.^[47]estimated that, based on the assumption that the annual output of carbon tetrachloride by-products equals the emission volume, the PCN emissions from carbon tetrachloride production in 2010 were 427 kg (536 g TEQ). Ba et al.^[82]estimated the PCN emissions from China's secondary nonferrous metals industry in 2007, showing that the total PCN emissions from this industry (in terms of toxic equivalent quantity, TEQ) were 1.27 g/a. Among these, secondary copper production was the primary emission source, with TEQ emissions of 0.86 g/a, accounting for 67.7% of the total emissions; secondary aluminum production had TEQ emissions of 0.39 g/a, accounting for 30.7%; and secondary lead and secondary zinc production had relatively low TEQ emissions of 0.009 g/a and 0.01 g/a, respectively, together accounting for 1.5%. Nie et al.^[84]estimated the annual PCN emissions from magnesium smelting in China to be 1,651 g. Liu et al.^[85]estimated the annual TEQ emissions of PCNs from iron ore sintering in China to be 1,390 mg. Li et al.^[94]estimated that in 2013, PCN emissions from iron ore sintering in China amounted to 568 kg (10.7 g TEQ), and in 2014, emissions totaled 574 kg (10.8 g TEQ). Dat et al.^[69]estimated the annual TEQ emissions of PCNs from 24 large municipal solid waste incineration plants in Taiwan, China, to be between 5.14 and 9.43 mg/a. Liu et al.^[83]estimated the annual TEQ emissions of PCNs from the global coking industry to be between 430 and 692 mg, with an average emission of 337 mg. Wang et al.^[99]estimated that between 2003 and 2015, sintering plants in China emitted a total of 17,364 g (in TEQ terms), of which 15,486 g were emitted via flue gases and 1,878 g via residues. From 2005 to 2015, the total PCN emissions from sintering plants showed a downward trend, with a cumulative reduction of approximately 1,549 kg (211 g TEQ). Yang et al.^[105]compiled an inventory of PCN emissions from waste incineration and metallurgical sources in China for 2014. The results indicated that the total PCN emissions in China in 2014 amounted to 511.6 kg (7,650.8 mg TEQ). Among these, waste incineration, secondary nonferrous metallurgy, electric arc furnace steelmaking, and iron ore sintering plants accounted for 38.8%, 15.4%, 29.2%, and 16.6% of the total emissions, respectively. The densely industrialized regions along China's eastern coast and Hebei in North China, characterized by high population density, are major sources of PCNs. Huang et al.^[106]estimated the provincial-level emissions of eight PCN congeners from 37 emission sources in mainland China between 1960 and 2019. The results showed that in 2019, total PCN emissions reached 757.0 kg, with Hebei Province ranking first among all provinces and the steel industry being the largest emission source.