Perfluorinated and Polyfluoroalkyl Compounds in the Atmosphere: A Review

Abstract

Perfluoroalkyl and polyfluoroalkyl substances (PFASs) are a class of synthetic organic compounds with extremely high chemical stability and environmental persistence that are widely used in the industrial sector and in consumer goods. Their strong C-F bonds make them difficult to degrade, meaning they can migrate through the atmosphere and settle over long distances, posing long-term risks to the global ecological environment and human health. This article systematically reviews the classification, physicochemical properties, concentration levels, spatial distribution, migration and transformation behaviors, and health and ecological impacts of PFASs in the atmosphere, along with related analytical detection techniques and pollution control methods. Studies show that short-chain PFASs are more likely to migrate through the atmosphere due to their high water solubility and volatility, while long-chain PFASs tend to be adsorbed onto particulate matter and display stronger bioaccumulation. Although atmospheric research on PFASs lags behind that focused on their dynamics in water and soil, the existing data still reveal a difference in their distribution and regional pollution characteristics in the gas and particle phases. Toxicological studies have confirmed that PFAS exposure is associated with liver injury, immunosuppression, developmental toxicity, and cancer risk and can threaten ecological security through the food chain. Currently, governance technologies are confronted with the challenges of low efficiency and high cost. In the future, it will be necessary to combine multi-media models, new analytical techniques, and international collaboration to promote the development of source control and innovative governance strategies.

1. Introduction

Perfluoroalkyl and polyfluoroalkyl substances (PFASs) are a class of synthetic fluorine-containing organic compounds that have been widely used in the industrial sector and in consumer goods (such as non-stick pan coatings, waterproof textiles, fire-fighting foams, etc.) since the middle of the 20th century. Their strong C-F bonds endow them with extremely high chemical and thermal stability but also make them difficult to degrade in the environment, meaning they are often referred to as “permanent chemicals”. In recent years, the problem presented by PFAS pollution has become increasingly serious worldwide, especially in industrial-intensive areas and remote regions where their presence has been detected. In addition to direct emissions into the atmosphere, PFASs in water bodies and the soil can also volatilize into the atmosphere in the form of vapor [1] because of their high mobility [2] and chemical stability [3], whether in a gaseous [4] or a granular state [4]. This means that they can migrate through the atmosphere and be transported to areas far from emission sources [5]. Epidemiological and toxicological studies have shown that PFAS exposure is associated with a variety of adverse health effects, such as liver and kidney diseases [6], immunotoxicity [7], and developmental toxicity [8], while the inhalation of PFASs can inhibit the function of pulmonary surfactants and cause acute toxicity [9]. Therefore, the atmosphere is the main site for the migration and transformation of PFASs, representing an important exposure route for their impact on human health [10]. It is therefore of great significance for understanding the dynamics of PFASs and the harm they pose to the environment, ecology, and human beings. Regarding PFASs in the atmosphere, previous studies have performed source distribution and health analysis of PFASs in water, soil, and the atmosphere, but there is a lack of systematic reviews focused on screening, evaluation, and control. The purpose of this article is to systematically review the classification and properties of PFASs in the tropospheric near-surface atmosphere; review the concentration levels and distribution characteristics of PFASs in gas and particle phases across different regions of China and compare them with findings from other countries; discuss the migration and transformation mechanisms that drive the persistence and long-range transport of PFASs in the atmosphere; evaluate toxicological evidence regarding the impacts of atmospheric PFASs on human, animal, and plant health; summarize the sampling and analytical techniques used for detecting atmospheric PFASs; review the current strategies for pollution control and identify their challenges; and finally provide a forward-looking perspective on research directions and policy implications for atmospheric PFAS management. By focusing on the situation in China and placing it within a global context, this paper seeks to highlight both the regional characteristics of atmospheric PFAS pollution and its broader implications for environmental health and regulatory control.

2. Classification and Properties of PFASs

Based on their functional groups and structures, PFASs (Figure 1) can be categorized into three main types: perfluoroalkyl substances, polyfluoroalkyl substances, and other derivatives.

According to carbon chain length, PFASs are divided into long-chain and short-chain compounds. Long-chain PFASs include perfluorocarboxylic acids (PFCAs) and perfluorosulfonic acids (PFSAs), with carbon chain lengths ≥ 7 (PFCAs) and ≥6 (PFSAs), such as PFOA (perfluorooctanoic acid, C8, PFCA) and PFOS (perfluorooctane sulfonic acid, C8, PFSA). Short-chain PFASs, developed as alternatives to long-chain compounds, have carbon chain lengths ≤ 6 (PFCAs) or ≤5 (PFSAs) and include PFBA (perfluorobutanoic acid, C4, PFCA), PFBS (perfluorobutane sulfonic acid, C4, PFSA), and GenX chemicals (hexafluoropropylene oxide dimer, C6, as a PFOA substitute). Some abbreviations related to FTOH substances include 1H,1H,2H,2H-perfluoro-1-octanol (6:2FTOH) and 1H,1H,2H, 2H-perfluoro-1-decanol (8:2FTOH).

The compounds listed in

Table 1. Common PFASs and their properties.

Classification	Substance	Chemical Properties	Physical Properties	Industrial Applications	Phase Tendency	Regulatory or Alternative Reasons
Short-chain PFASs	Perfluorohexanoic acid (PFHxA)	Shorter carbon chains slightly improve degradability compared to long-chain PFCAs	b.p. ~155 °C; m.p. ~30 °C, high water solubility	Used in textiles, paper cleaners, and surface treatments	Mixed (gas/particle partitioning varies)	Still in current use, considered less bioaccumulative but more mobile in the environment
	Perfluorobutanoic acid (PFBA)	Ultra-short chains enhance mobility and degradability	b.p. ~120 °C; m.p. ~20 °C, high water solubility	Used in pesticides, pharmaceutical intermediates, and electronic industry etchants
	Perfluorobutane sulfonic acid (PFBS)	Short chains improve degradability but retain high stability	b.p. ~160 °C; m.p. ~50 °C, moderate water solubility	Used in electronic etching, firefighting foams, cosmetics, and cleaners
	GenX (hexafluoropropylene oxide dimer acid)	Designed as a PFOA substitute with lower bioaccumulation but high stability	b.p. ~200 °C; m.p. ~60 °C, high water solubility	Used in fluoropolymer production, waterproof materials, and high-performance coatings
Long-chain PFASs	Perfluorooctanoic acid (PFOA)	Extremely stable, resistant to degradation, high surface activity	b.p. ~189 °C; m.p. ~40–50 °C, moderate water solubility	Formerly used in polytetrafluoroethylene emulsifiers, waterproof coatings, firefighting foams, and semiconductor manufacturing	Particle-phase prone (very low vapor pressure)	Listed under Stockholm Convention; replaced by C4–C6 PFCAs or fluorine-free surfactants
	Perfluorononanoic acid (PFNA)	High chemical stability, resistant to heat and chemical corrosion	b.p. ~220 °C; m.p. ~60 °C, low water solubility	Used in fluoropolymer production, waterproof coatings, and firefighting foams
	Perfluorooctane sulfonic acid (PFOS)	Exceptional chemical and thermal stability, resistant to degradation	b.p. ~280 °C; m.p. ~90 °C, low water solubility	Formerly used in firefighting foams, metal plating antifogging agents, textile/leather waterproofing, and photoresists in electronics	Particle-phase prone	Restricted by Stockholm Convention due to persistence and toxicity; replaced by PFHxS (short-chain) or fluorine-free foams
	Perfluorohexane sulfonic acid (PFHxS)	Shorter chains slightly improve degradability compared to PFOS	b.p. ~210 °C; m.p. ~70 °C, moderate water solubility	Used in firefighting foams and chrome plating processes
	ADONA (ammonium perfluoro-3,6-dioxaheptanoate)	Designed as a PFOA substitute with lower bioaccumulation and environmental persistence	b.p. ~220 °C; m.p. ~70 °C, high water solubility	Applied in firefighting foams and industrial fluoropolymer production	Particle-phase prone	The European Union has listed it as a substance of very high concern (SVHC) and it is currently under review by REACH
	Fluorotelomer alcohols (FTOHs)	Precursor compounds that readily transform into other PFASs, highly volatile	b.p. 100–150 °C; m.p. −20 °C, low water solubility	Used in waterproof and oil-resistant treatments for textiles and paper	Gas-phase prone (high vapor pressure, semi-volatile)	Many textile/apparel applications phased out due to degradation to PFOA/PFOS; replaced by C6-based FTOHs or non-fluorinated coatings
	Perfluorooctane sulfonamide (FOSA)	PFOS precursor with high chemical stability	b.p. ~250 °C; m.p. ~80 °C, low water solubility	Intermediate for pesticides and waterproofing agents	Gas-phase prone	Widely used in the waterproofing of pesticides, textiles, and paper products, but is listed as a restricted substance by the Stockholm Convention due to its degradability into PFOS
	Perfluorooctane sulfonamide ethanol (FOSE)	Strong stability, mainly converted into PFOS	b.p. ~260 °C; m.p. ~90 °C, low water solubility	Used in coatings, polymer processing, and stain-resistant carpet treatments	Gas-phase prone	Because it is the main precursor of PFOS, it has been largely phased out; subject to the Stockholm Convention.
	Perfluoroalkyl phosphates (PAPs)	High chemical stability, mainly converted into PFOA/PFOS	b.p. ~300 °C; m.p. ~100 °C, low water solubility	Oil-resistant coatings for food packaging; cosmetics and personal care products	Mostly present in the granular phase, but the gas–granular biphasic distribution can be detected in some medium-chain PAPs.	Can be degraded into PFOA, PFHxA, etc.; both the European Union and the United States have imposed restrictions on its use in food packaging
	Perfluoropolyethers (PFPEs)	High-performance lubricants and coolants with exceptional chemical and thermal stability	b.p. ~350 °C; m.p. ~120 °C, very low water solubility	High-temperature lubricants; coolants in semiconductor industry; sealing materials	Particle-phase prone	Some varieties have been given attention because their degradation products may form PFOA; some PFPEs in the European Union have been included in the REACH restriction list

are difficult to degrade due to the high stability of their C-F bonds, meaning that they can be widely diffused through atmospheric transport, the water cycle, and other pathways, posing long-term risks to ecosystems and human health. The table clearly shows that PFASs generally have extremely high chemical and thermal stability, enabling them to persist in the environment for a long time [11]. Their physical properties, such as their boiling and melting points, are relatively high, while their water solubility varies due to differences in chain length and structure [12]. However, the melting and boiling points of different types of PFASs vary significantly. For instance, experimental data from the new alternative GenX (HFPO-DA) indicate that its melting point is approximately −98 °C, while its boiling point is about 64.7 °C [13]. These values were directly obtained from the Chemical Safety Data Sheet (SDS) and can provide reliable physicochemical parameter support for subsequent research. In contrast, the boiling points of traditional long-chain PFASs (such as PFOA and PFOS) are generally above 150 °C, while that of some compounds is difficult to determine due to their extremely high thermal stability [14]. Therefore, it can be considered that GenX is more prone to migration in terms of physical and chemical properties, while other long-chain PFASs exhibit stronger environmental persistence. From the classification provided in the table, we can see that short-chain PFASs usually have higher water solubility and lower bioaccumulation, while long-chain PFASs are more persistent and more easily accumulate in organisms [15].

3. Concentration Levels and Behavior of PFASs in the Atmosphere

3.1. Concentration Levels of Atmospheric PFASs

Up to now, research on the environmental pollution characteristics of PFASs in China and other countries has mainly focused on water bodies and soil, while studies on PFASs in the atmosphere are relatively scarce. The atmosphere is vital for human survival, and atmospheric transport is also one of the important pathways for PFAS transport. The distribution of PFASs (

Table 2. The concentration levels of PFASs in the atmosphere, reported in the literature from China and other countries.

Sampling Location	Sampling Time	Detected PFASs Compounds and Their Concentration Levels		Unit	References
		Gas Phase	Particle Phase
Northern coast of Germany (Büsum)	2011.8–2012.10	6:2FTOH: 0.7–40	6:2FTOH: <0.1–0.5	pg·m−3	[17]
		8:2FTOH: 4.8–85	8:2FTOH: <0.2–0.8
		10:2FTOH: 0.8–17	10:2FTOH: <0.1–0.4
		12:2FTOH: 0.4–5.4	12:2 FTOH: <0.1–0.6
		6:2FTA: <0.2–1.2	6:2FTA: <0.2–0.2
		8:2FTA: 0.2–2.4	8:2FTA: <0.1–0.2
		MeFBSA: 0.3–12	MeFBSA: <0.1–0.2
		MeFOSA: 0.1–1.9	MeFOSA: <0.1–0.1
		EtFOSA: <0.1–1.1	EtFOSA: <0.1–0.1
		MeFBSE: 0.3–2.6	MeFBSE: <0.1–0.4
		MeFOSE: 0.2–1.4	MeFOSE: <0.2–0.2
		EtFOSE: <0.1–0.9	EtFOSE: <0.1–0.3
Northern South China Sea	2013.9–10	6:2FTOH: 0.7–8.6	6:2FTOH: <MDLs (0.1)	pg·m−3	[18]
		8:2FTOH: 12.5–75.8	8:2FTOH: <MDLs (2.1)
		10:2FTOH: 3.3–17.8	10:2FTOH: <MDLs (0.1)
		12:2FTOH: 1.3–6.0	12:2 FTOH: <MDLs (0.1)
		FTAs: <0.1–0.4	FTAs: <MDLs (0.1)
		FASA and FASEs: 0.1–3.6	FASA and FASEs: <MDLs (0.1)
Košetice, Czech Republic	2012.1–2014.12	∑PFCAs: 0.03–2.08		pg·m−3	[19]
		∑PFSAs: 0.02–0.85
		∑FOSA/Es: ND-0.18
Ny-Alesund in the Arctic region	2011.9–2012.9	6:2FTOH: 3.6–24	FOSAs (MeFBSA, MeFOSA, EtFOSA): >FOSEs (MeFBSE, MeFOSE, EtFOSE)	pg·m−3	[20]
		8:2FTOH: 4.7–28.6
		10:2FTOH: 5.8–33.7
		12:2FTOH: 5.4–37.4
		6:2FTA: <0.1–0.1
		8:2FTA: <0.1–0.5
		MeFBSA: 0.3–2.6
		MeFOSA: 0.1–0.9
		EtFOSA: 0.2–1.2
		MeFBSE: 0.4–3.5
		MeFOSE: 0.4–3.2
		EtFOSE: 0.4–3.1
Sweden, Finland, Italy, Norway, Estonia, and other places	2007–2010	6:2FTOH: 0.834–997	PFHxA: Personal sampling (0.7–478) Regional sampling (1.39–333)	μg·m−3	[21]
		8:2FTOH: 0.001–42.7
		10:2FTOH: 0.002–1.96	PFOA: Personal sampling (1.2–47) Regional sampling (2.11–52.8)
		PFHxA: 0.027–14.9
		PFOA: 0.002–4.89	PFDA: Personal sampling (0.28–54.8)
		PFCA: 0.01–6.59	PFTDA: Personal sampling (0.33–279)
Beijing	2018.11–2019.7 2021.6 and 12	∑PFASs: 130.84 ± 92.01 ∑PFCs: 9.33	∑PFASs:35.91 ± 17.28 PFOA: 4.67 PFOS: 3.28 PFBA: 1.22	pg·m−3	[22,23]
Xiamen	2018.6–2019.5	PFBA: 2.4 6:2Cl-PFESA: 0.48 PFHxA: 0.46 PFOA: small amount PFOS: extremely rare	6: 2Cl-PFesa: 2.1 PFHxA: not specified PFOA: accounts for 18% PFOS: accounts for 45%	pg·m−3	[24]

) in the gas and particle phases affects their form and deposition mode in the atmosphere. However, compared with water bodies and sediments, research on monitoring PFASs in the atmosphere is still very limited. On the one hand, most PFASs have low volatility, and only a few precursor compounds such as FTOHs and sulfonamide ethanol can exist in the atmosphere in a gaseous or granular form, thereby limiting the number of target species that can be monitored. On the other hand, the concentration of PFASs (Figure 2) in the atmosphere is usually at the pg/m³ level, posing extremely high requirements for sampling and detection sensitivity. In addition, PFASs are prone to complex photochemical transformations in the atmosphere, generating multiple intermediates and stable end products, which increase the uncertainty of monitoring and traceability. Although there is already evidence indicating that long-distance atmospheric transmission is an important pathway for the global distribution of PFASs, the current related monitoring data still display fragmented and unsystematic characteristics, causing atmospheric research to lag significantly behind that focused on other environmental media [16]. In previous atmospheric research on PFASs, reports from outside China have mainly focused on developed coastal countries such as northwest Europe, the United Kingdom, Germany, and the United States. However, this research has been ongoing for over a decade, lacking long-term and sustained monitoring and detection. Research on PFASs in the atmosphere in China has accelerated in the past decade, but related studies remain limited, with the main research sites including coastal cities such as Tianjin, Yantai, and Xiamen, as well as some areas in the South China Sea or near pollution sources such as factories.

On a global scale, previous studies have performed sampling from 2010 to 2021, and the spatial scope includes both China and international regions, such as Europe, and domestic areas of China like Beijing, the South China Sea, and remote regions like the Arctic. China has carried out more than twenty atmospheric PFAS monitoring projects, with studies related to the granular phase being predominant, as the sampling techniques used (such as solid-phase adsorbents and high-volume air sampling) are mainly applicable to the analysis of PFASs in particulate matter, while less attention has been given to gas-phase PFASs [25]. In a study focused on Xiamen, the concentration of particulate PFASs was significant (approximately 4.11–67.41 pg/m³, with an average of 26.56 pg/m³), highlighting the status of particulate matter as the main carrier of atmospheric PFASs [26]. Multi-phase sampling in Shijiazhuang and a structural gas/particle distribution study in Beijing further confirmed that most PFASs tend to be adsorbed onto particulate matter (distribution coefficient < 0), while the specific gravity of the gas phase is relatively low [27,28]. FTOHs in the northern South China Sea are the most significant group of PFASs, accounting for 95.2% to 99.3% of all PFASs. The low concentrations of short-lived PFASs, such as FTAs and FASEs, indicate that their sources may be related to long-distance transmission. The Asian continent is an important source of PFASs in the atmosphere of the northern South China Sea, with the concentration of PFASs being significantly higher in samples affected by continental airflows than in samples not subject to this effect. The concentration of PFASs in the atmospheric particle phase from the northern South China Sea is extremely low, further confirming that they mainly exist in a gaseous form in the atmosphere.

In a study [29] conducted in the southwestern suburbs of Chengdu, China, Zhang Yun et al. studied the concentration level of PFASs in atmospheric particulate matter during winter. Their research results showed that among the 25 types of PFASs identified, a total of 10 were found in this area, with the total concentration (∑10PFASs) ranging from 4.58 to 647.59 pg·m⁻³ and the average value being 140.81 pg·m⁻³. Among these PFASs, the concentration of PFBA was the highest, with an average of 133.18 pg·m⁻³, contributing 69% of the total. Next was PFOA, with an average concentration of 2.98 pg·m⁻³. In contrast, the concentrations of PFHpA and 6:2 FTSA were relatively low, at 1.36 pg·m⁻³ and 1.22 pg·m⁻³, respectively. These concentrations were comparable to the PFAS levels in Beijing (157 pg·m⁻³) and Zhejiang [30] (251.93 pg·m⁻³), although the concentration of PFBA was significantly higher than that in Beijing (1.22 pg·m⁻³) and Xiamen (2.4 pg·m⁻³). This research also found that the concentration of short-chain PFASs (such as PFBA) was relatively high, which might be related to the use of substitutes for long-chain PFASs and local industrial emissions. In the research conducted by North China Electric Power University in Beijing, PFOA contributed the most to the overall PFAS concentration, accounting for 37.4% in the gas phase and 36.7% in the particulate phase, followed by PFHpA, which accounted for 36.4% in the gas phase and 24.7% in the granular phase.

Research on the gas-phase concentration levels of PFASs is relatively scarce in some countries, with most studies focusing on the contamination characteristics of PFASs in the particulate phase. It is worth noting that although the concentration levels in New Jersey and the Arctic region differ significantly, the underlying source mechanisms complement each other: the former represents a typical industrial and urban emission source area, while the latter represents a distant subsidence zone of long-distance atmospheric transmission. Combining the concentration characteristics of these two regions for discussion can not only help us to better explain the regional distribution differences in PFASs, but also highlight their global migration potential. In recent research [31] conducted in the Americas, Ying Yao et al. found that in the urban atmosphere of northern New Jersey, USA, the concentration of PFASs in the gas phase (197.7 ± 47.9 pg·m⁻³) was higher than that in the particle phase (48.3 ± 47.9 pg·m⁻³), indicating a higher risk to human health from inhalation. Short-chain alternatives to PFASs, such as perfluoroheptanoic acid, were found to have a concentration in the gas phase (142.6 ± 28.0 pg·m⁻³) that was higher than the maximum PFAS concentration level set out by the US EPA’s regulations. Research on the distribution of PFAS gas particles was also carried out in the Czech Republic, and it was found that short-chain PFCAs (PFPeA, PFHxA) are distributed in both the gas phase and the particle phase, with a relatively low particle-phase fraction (θ), averaging 0.30–0.65. Long-chain PFCAs (PFDA, PFUnDA) tend to exist in the granular phase, with θ increasing with the length of the carbon chain, and the maximum θ can reach 0.65–0.74. The average θ of PFOA is relatively low, at 0.14 ± 0.33, PFOS mainly exists in the particulate phase, with θ = 0.80 ± 0.29, and PFDS mainly exists in the gas phase, with θ = 0.12 ± 0.29. Other PFSAs (PFBS, PFHxS) are relatively evenly distributed in the two phases, with θ = 0.38–0.65. FOSA/Es mainly exist in the gas phase, with θ = 0.03–0.22, and among them, the θ of FOSA is the lowest, θ = 0.03 ± 0.09, while that of EtFOSE is slightly higher, θ = 0.22 ± 0.38. It can be seen from this that the distribution behavior of PFASs in the gas phase and the granular phase varies depending on the compound type and chain length. Short-chain PFCAs and FOSA/Es tend to exist in the gas phase, while long-chain PFCAs and PFOS tend to exist in the granular phase. In research conducted on the Arctic region, the PFASs in the atmosphere were mainly found to be in the gaseous phase (91%), with 8:2 FTOH being the main component. The concentration of FOSAs in the particle phase is relatively high, while the concentration of FTA is relatively low due to its anion characteristics. In research conducted in the regions of Northern Europe where skiing is prevalent, the particle phase distribution of PFASs was divided into two parts—among the inhalable particles (<100 μm), PFHxA accounted for 45%, PFTDA for 16%, and PFOA for 14%, while among the lung-penetrable particles (<4 μm), PFHxA accounted for 28%, PFTDA for 27%, and PFOA for 16%. The gas phase mainly consists of FTOHs, especially 8:2 FTOH, which almost entirely exists in the gas phase and has not been detected in the particle phase. PFHxA and PFOA are dominant in the particle phase, with relatively high concentrations, and are found to be carbon chain homologs. The data obtained from this research indicate that the concentration of PFASs that ski waxing technicians are exposed to (at the μg·m⁻³ level) is significantly higher than that for the general population (at the ng·m⁻³ level) by an order of magnitude. The above data indicate that ski waxing technicians are exposed to high concentrations of PFASs in both the gaseous and granular phases, especially 8:2 FTOH and PFCAs, posing a significant risk to occupational health. The following are some abbreviations that appear in the table: MeFOSE: 2-(N-methylperfluoro-1-octanesulfonamido)-ethanol; EtFOSA: N-ethyl-perfluoro-octane Sulfonamide; MDL: method detection limit. All concentrations have been converted to pg·m⁻³ for comparability (1 μg·m⁻³ = 1 × 10⁶ pg·m⁻³; 1 ng·m⁻³ = 1 × 10³ pg·m⁻³).

3.2. Migration and Transformation Processes of PFASs in the Atmosphere and Their Persistence in the Atmosphere

PFASs mainly enter the atmosphere through industrial emissions and volatilization from consumer products and can exist in a gaseous or particulate form. Among them, short-chain substances with strong volatility can migrate over long distances, while long-chain compounds are mostly adsorbed on particulate matter. PFASs spread globally through atmospheric circulation and condense and settle in cold regions under the influence of temperature. Eventually, they return to the surface through dry or wet deposition, and some may be released a second time. During this migration process, PFASs may undergo photodegradation or oxidative transformation. Their distribution is jointly regulated by the characteristics of the compound and meteorological conditions, leading to persistent pollution and even accumulation in remote areas.

3.2.1. Migration Process of PFASs in the Atmosphere

The long-distance atmospheric transport of PFASs is an important factor influencing their distribution in a region. Many PFASs, such as FTOHs and PFCAs, have relatively high vapor pressures and can exist in gaseous form in the atmosphere and be transported over long distances. For example, the atmospheric retention time of 8:2 FTOH can reach 80 days, which is sufficient for its global spread. The global production of 8:2 FTOH is approximately 5 × 10⁶ kg per year, of which 40% is produced in North America [32]. In the abovementioned study on the Arctic, the average ratio of 8:2, 10:2, 6:2, and 12:2 FTOHs in the atmosphere was 4.8:0.9:1.0:0.4. This ratio was higher than that in urban areas (Toronto: 1.1:0.2:1.0), indicating that FTOHs mainly reached the Arctic through long-distance (Figure 3) atmospheric transport, demonstrating that this is a key factor affecting the distribution of PFASs in the Arctic [20]. Recent research in a factory in the Netherlands [33] showed that PFASs such as HFPO-DA (GenX) can be transported over long distances through the atmosphere, with model predictions showing that HFPO-DA can still be detected thousands of kilometers away from the Dordrecht factory—for example, the annual deposition in Reykjavik, Iceland, is 0.5–2.4 ng·m⁻²yr⁻¹. This research also found that wind direction is one of the influencing factors in PFAS concentration—when the wind blows from the direction of the factory, the PFAS concentration in the air increases significantly.

The dry deposition of PFASs is one of the most important ways in which they are removed from the atmosphere and enter soil and water bodies. In the abovementioned study on a Dutch factory, during the deposition process, PFASs settled from the atmosphere onto the surface through both dry and wet deposition. The model showed that the daily deposition amount of HFPO-DA in the area close to the factory (within 556 m) reached 1 µg·m⁻²yr⁻¹, gradually decreasing as the distance increases. There are significant differences in the dry deposition rates of PFASs. Long-chain PFCAs and PFSAs, due to their extremely low vapor pressure and strong hydrophobicity, mainly adsorb on fine atmospheric particulate matter, and their dry deposition rates are relatively low, usually within the range of 0.1–0.5 cm·s⁻¹ [20]. Short-chain PFCAs have a slightly higher dry deposition rate due to their higher water solubility and the fact that some of them exist in a gaseous form. Although volatile precursors mainly exist in a gaseous form, their dry deposition rates are controlled by the exchange interface process with the surface (vegetation, water bodies, etc.), and the range of the variation rate is relatively large (~0.1–1.0 cm·s⁻¹) [34]. PFASs mainly exist in gaseous and particulate states in the atmosphere, and their distribution between these states is a dynamic equilibrium process. The gas/particle distribution of PFASs in the atmosphere is influenced by factors such as temperature, relative humidity, and the nature of the particulate matter. Under conditions of low temperature or high humidity, PFASs are more inclined to adsorb onto the surface of particulate matter and then settle through gravity, as well as directly settling from the gaseous state onto the ground, thereby increasing their deposition rate in the atmosphere. Generally speaking, shorter chain PFASs (6:2 FTOH) are more likely to exist in a gaseous form, while longer chain compounds (PFOA) tend to adsorb onto particulate matter [20]. The dry deposition rate of PFASs adsorbed on particulate matter is affected by its particle size, density, and atmospheric turbulence. Larger particulate matter (such as PM₁₀) settles faster, while smaller particulate matter like PM_2.5 may remain suspended in the atmosphere for a longer period of time. Wet deposition is also an important way in which PFASs are removed from the atmosphere, mainly including deposition onto the ground through precipitation (rain, snow, and fog) [35]. The wet deposition process is influenced by the intensity of precipitation, the concentration of PFASs in the atmosphere, and the nature of aerosols. A previous study showed that the median concentration range of PFASs after wet deposition was 2.4 to 4.5 ng·L⁻¹, and there was no significant difference in concentration among different sites, indicating that PFASs, especially ionic PFASs, are widely distributed in the atmosphere [36].

3.2.2. The Transformation Process of PFASs in the Atmosphere

PFASs undergo photochemical transformation in the atmosphere. Although PFASs have extremely high chemical stability, their environmental behavior is still significantly affected by their chain length—for instance, some short-chain PFASs containing fluorine, such as perfluorobutyric acid, may undergo photolysis under strong ultraviolet light, generating low-fluorinated organic acids or other small-molecule compounds. However, the efficiency of this transformation process is very low, mainly involving short-chain PFASs, while long-chain PFASs like perfluorooctanoic acid rarely undergo photolysis [35]. The current research consensus indicates that the reaction with ·OH radicals in the atmosphere represents the most important degradation pathway for neutral PFAS precursors [37]. Their atmospheric lifetime (τ) can be estimated according to the formula *τ = 1/(k·OH·[OH])*, where k·OH is the second-order rate constant of the reaction with ·OH, and [·OH] is the average concentration of atmospheric ·OH radicals (usually taken as 1.0 × 10⁶ molecules·cm⁻³). Based on kinetic parameters determined in the laboratory, the estimated atmospheric half-lives of the main precursors are as follows:

Fluorinated alcohol (FTOHs): The k·OH of the reactions between 6:2 FTOH, 8:2 FTOH, and 10:2 FTOH and ·OH is approximately (1.0–1.5) × 10⁻¹² cm³·molecule⁻¹·s⁻¹ [38]. Based on this, their atmospheric half-life is estimated to be approximately 10 to 20 days. This is in the same order of magnitude as those inferred from field studies for 6:2 FTOH (~50 days), 8:2 FTOH (~80 days), and 10:2 FTOH (~70 days) [39], verifying that ·OH oxidation is the dominant degradation mechanism.

Sulfonamides (FOSA/FOSE): Taking N-methylperfluorooctane sulfonamide ethanol (MeFOSE) as an example, its k·OH is relatively high, at approximately 5.8 × 10⁻¹² cm³·molecule⁻¹·s⁻¹ [40], and thus its atmospheric half-life is shorter, at about 2 to 3 days.

Straight-chain PFCAs/PFSAs: The direct photolysis or reaction of PFOA, PFOS, and other ionic substances with ·OH is extremely slow, meaning they have an atmospheric half-life of tens of days or even longer. Their final removal is therefore more dependent on physical deposition processes rather than chemical transformation [41].

In summary, short-lived precursors can undergo significant transformation during regional-scale transport, while long-lived precursors and end products have the ability to undergo global-scale migration. This explains why these substances can also be detected in remote polar environments.

In addition, PFASs undergo atmospheric chemical reactions in the atmosphere. The core structure of PFASs is the perfluoroalkyl chain (CF₃[CF₂]ⁿ⁻), which has extremely high chemical stability, meaning that they are not easily biodegraded or chemically degraded in the natural environment. Some studies have indicated that the chemical bond energy of PFASs such as PFOA and PFOS is extremely high, making them almost non-degradable under natural conditions [35]. Under specific chemical conditions, PFASs may undergo limited chemical transformations. For instance, PFASs may react with hydroxyl radicals (OH·) in the atmosphere to form fluorine-containing carboxylic acids or other intermediate products [32]. However, the rate of this reaction is very slow, and the degree of transformation is limited. Short-chain PFASs have a higher vapor pressure and are more likely to volatilize into the atmosphere, thus having a stronger transport potential, reflecting the trend of short-chain PFASs replacing long-chain PFASs [42]. In addition, chemical reactions on the surface of particulate matter may also lead to the partial transformation of PFASs, for instance by undergoing oxidation reactions with other pollutants such as ozone and free radicals to form more degradable compounds [43].

PFASs also undergo biotransformation in the atmosphere. Although they rarely undergo biodegradation in the atmosphere because their chemical structure lacks the active sites required for microbial degradation, some studies [44] have shown that certain microorganisms may carry out a limited transformation of PFASs under specific conditions. For instance, in the literature, researchers used a mixed microbial culture (acclimated with ethanol as the carbon source) to conduct a degradation experiment on 8:2 FTOH under aerobic conditions, while setting up a sterile control to eliminate the influence of non-biodegradable factors. However, this transformation process is very rare in the atmospheric environment due to a lack of microorganisms and suitable environmental conditions [45]. PFASs have high hydrophobicity and oleophobicity and can combine with proteins and lipids in organisms, leading to their accumulation within the body. This bioaccumulation further increases the persistence of PFASs in the environment, as they can be transferred through the food chain and accumulate in high-trophic organisms [42].

3.2.3. Limitations of Laboratory Studies Versus Real Atmospheric Conditions

Although numerous studies have investigated the migration, transformation, and removal of PFASs under laboratory conditions, it is important to note that these results may not fully reflect their actual behavior in the atmosphere. Laboratory experiments are often carried out under controlled settings, where factors such as temperature, humidity, radiation, and the presence of co-pollutants are simplified or isolated. In contrast, the real atmosphere is a far more complex system, with dynamic meteorological conditions, diverse aerosol compositions, and multiple competing chemical and physical processes. These environmental variables can significantly alter the persistence, partitioning, and degradation efficiency of PFASs. Therefore, while laboratory results are valuable in elucidating potential mechanisms, their direct extrapolation to atmospheric scenarios should be treated with caution. Future research should strengthen the integration of laboratory experiments, field monitoring, and atmospheric modeling to bridge this gap and provide more realistic assessments of PFAS behavior and risks in the environment.

4. Toxicity of PFASs in the Atmosphere

4.1. Hazards of PFASs in the Atmosphere to Human Health and Assessment Methods

PFASs in the atmospheric environment not only enter the human body through the respiratory pathway but have also been found to have significant systemic and reproductive development toxicity effects, even at relatively low concentrations (

Table 3. Summary of effects.

Compound	Exposure Pathways/Scenarios	Concentration Threshold/Level	Category of Health Effects	Key Points	References
PFOA	Inhalation (ambient air, adult 70 kg, 20 m3/day assumption)	≈10,500 pg·m−3 (inversely deduced from the ATSDR minimum risk level MRL = 3 × 10−6 mg·kg−1·d−1)	Comprehensive (rRisk assessment)	Most areas are far below this range, but the pollution sources may be close to or exceed it	[49]
	Inhalation (animal experiments)	LC50 = 980 ng·m−3 (acute inhalation toxicity)	Comprehensive/systemic toxicity	Systemic damage	[50]
PFOS	Inhalation (animal experiments)	LC50 = 5.2 mg·L−1 (acute inhalation toxicity)	Comprehensive/systemic toxicity	Respiratory tract and liver damage	[50]
PFOA; PFOS	Internal load (thyroid tissue)	The median PFOA ≈ 2 ng·g−1; the median PFOS is approximately 5.3 ng·g−1	Endocrine (thyroid)	Related to thyroid hormone disorders	[51]
6:2 Cl-PFESA; PFOA	Cerebrospinal fluid (CSF)	CSF: 6:2 Cl-PFESA ≈ 0.051 ng·mL−1; PFOA ≈ 0.078 ng·mL−1; permeability: 0.84%, 1.07%	Endocrinology/metabolism and neurology	Activate PPAR, and there is a positive correlation with blood glucose	[51]
Multiple PFASs (PFOA as an example)	Serum inflammatory markers	CRP per logarithmic unit ↑ → R ↑ 0.109 of PFOA	Immunity/inflammation	CRP is positively correlated with permeability	[51]
PFBA	Human tissue	Lung ≈ 807 ng·g−1; kidney ≈ 263 ng·g−1	Development/system	High accumulation of short-chain PFAS in the lungs and kidneys	[51]
PFHxA	Human tissue	Liver ≈ 68.3 ng·g−1; brain ≈ 141 ng·g−1	Neurology and metabolism	Higher concentration in the brain and liver	[51]
PFOA (and several PFASs)	Long-term exposure	No threshold has been established; epidemiological associations (renal cancer, testicular cancer, etc.)	Cancer	Associated with an increased risk of multiple types of cancer	[52]
Multiple PFASs→HCC	Mechanism inference	Molecular/transcriptome evidence; 6 core genes, binding energy −5.3 to −7.6 kcal·mol−1	Cancer (HCC)	Mechanism chain support	[53]
FTOHs (8:2, etc.)	Indoor air intake	250–82,300 pg·m−3; adult inhalation: 1.04–14.1 ng·kg−1·d−1	Exposure level (comparison)	Indoor air intake is much higher than dust	[42]
Multiple PFASs	Occupational exposure (skiing waxing)	In the order of μg·m−3	Exposure level (occupation)	Occupational exposure has significantly increased	[21]

“↑" indicates an increase.

). In [46], the FTOH level in indoor air reached 250–82,300 pg/m³ and the adult inhalation intake was 1.04–14.1 ng/kg·day, which was significantly higher than the dust exposure level, highlighting the importance of air inhalation. In animal experiments, the inhaled LC₅₀ of PFOS was 5.2 mg/L and that of PFOA was 980 ng/m³, both of which led to systemic damage to the respiratory tract, liver, etc. [47]. Most crucially, atmospheric exposure to PFASs has been closely associated with thyroid dysfunction, immunosuppression, decreased sperm quality, and fetal developmental toxicity [48], supporting the view that it is a key pollutant with both environmental and health risks.

PFASs are carcinogenic, with epidemiological studies [54] showing that PFOA exposure is associated with an increased risk of testicular cancer, kidney cancer, non-Hodgkin’s lymphoma, etc. Recently [55], Hong et al. systematically analyzed the effects of PFASs on the occurrence and development of hepatocellular carcinoma (HCC) and their potential molecular mechanisms. In this study, these authors analyzed transcriptome data from multiple public databases, including core targets TCGA-LIHC (n = 355), LIRI (n = 202), LICA (n = 152), CHCC (n = 159), GSE14520 (n = 221), and GSE54236 (n = 78) (where “n” represents the sample size and serves as the foundation of the research), to identify differentially expressed genes (DEGs) associated with PFAS exposure and HCC. Through differential gene expression analysis, the researchers identified 174 target genes that are commonly associated with PFAS and HCC. These genes are significantly enriched in metabolic signaling pathways and are involved in processes such as lipid metabolism, glucose metabolism, and drug metabolism. Among these target genes, the researchers further screened out six core target genes (APOA1, ESR1, IGF1, PPARGC1A, SERPINE1 and PON1) by integrating machine learning and protein–protein interaction (PPI) network analysis. These genes may play an important role in the progression and prognosis of HCC. For instance, the expression level of APOA1 in normal tissues is significantly higher than that in tumor tissues, and it is associated with the overall survival (OS) of HCC patients. Furthermore, as verified by RT-qPCR and immunohistochemistry (IHC), the expression levels of these core target genes in tumor tissues were significantly lower than those in normal tissues (p < 0.05). The researchers also developed a survival risk model for PFAS-related HCC (PFASRHSig), which integrates 10 machine learning algorithms and 101 algorithm combinations to select 14 genes from 41 genes with prognostic value to construct a risk score. In the TCGA-LIHC training dataset and five validation datasets, the overall survival of patients in the high-risk group was significantly lower than that in the low-risk group (p < 0.001). The diagnostic accuracy of this model was evaluated by means of ROC-AUC analysis. AUC stands for area under the ROC curve, indicating the accuracy of the model’s prediction, and is a core metric for evaluating the performance of machine learning models. The AUC values of survival rates at 1 year, 3 years, and 5 years were all relatively high in the different cohorts. For example, they were 0.728, 0.708, and 0.732, respectively, in the TCGA-LIHC cohort. In addition, through molecular docking simulation, the researchers found that PFAS compounds have a strong binding affinity with these six core target proteins, with binding energies ranging from −5.3 to −7.6 kcal·mol⁻¹. This indicates that these interactions are spontaneous and may play an important role in the occurrence of PFAS-induced liver cancer. The results of the analysis of the molecular pathways related to hepatocellular carcinoma (HCC) obtained through bioinformatics modeling indicated that PFAS exposure might induce abnormal expression associated with oxidative stress, inflammatory response, and the PPAR signaling pathway. Although these findings mainly originated from the analysis of cell and transcriptome databases and the dose levels were higher than the actual concentrations in the atmospheric environment, the findings remain important. Based on existing epidemiological evidence, it can be speculated that long-term, low-dose exposure to atmospheric PFASs leads to their continuous accumulation in the human body and may increase the risk of HCC through a similar molecular mechanism. Therefore, bioinformatics modeling provides a molecular-level explanation for the potential health effects of low-level atmospheric exposure and helps to establish a mechanism chain from atmospheric concentration levels to disease risks.

PFASs can interfere with the endocrine system. PFASs such as PFOS and PFOA may disrupt the function of thyroid hormones, affecting metabolism, growth, and development. Some studies [51] have shown that PFASs are associated with abnormal thyroid hormone levels, especially in pregnant women and children, who are more susceptible. PFOA, PFOS, and 6:2 Cl-PFESA can cross the blood–cerebrospinal fluid barrier (BCSFB), and although the permeability is relatively low (e.g., 0.84% for 6:2 Cl-PFESA and 1.07% for PFOA), their persistent presence in the cerebrospinal fluid indicates that they have the ability to enter the central nervous system [56]. Among these, specific data have also demonstrated that PFOA and PFOS were detected in human thyroid tissues, with median concentrations of 2 ng·g⁻¹ and 5.3 ng·g⁻¹, respectively [57]. PFASs may also interfere with lipid and glucose metabolism by activating the peroxisome proliferator-activated receptor (PPAR) pathway, increasing the risk of diabetes. It was observed that PFOA, PFOS, and their substitute 6:2 Cl-PFESA were all found at high levels in the serum and could cross the blood–cerebrospinal fluid barrier to varying degrees, thus entering the central nervous system. Among these compounds, the ability of 6:2 Cl-PFESA to activate the PPAR signaling pathway was significantly higher than that of traditional PFASs, showing a stronger potential for metabolic interference. The authors of [53] point out that the biological activity of such new substitutes may be even higher than that of the “previous generation” of pollutants which have been phased out, suggesting that their potential health risks cannot be ignored. Further clinical analysis showed that the cerebrospinal fluid permeability of PFASs was significantly positively correlated with the serum glucose concentration. In particular, the R of PFOA and 6:2 Cl-PFESA with the glucose level was β = 0.320 (95% CI: 0.108, 0.532) and β = 0.392 (95% CI: 0.130, 0.654), respectively, indicating a positive linear relationship. This discovery supports the idea that PFASs may be involved in the occurrence mechanism of metabolic diseases such as diabetes by influencing glucose metabolism [56].

PFASs can also suppress the immune system. PFAS exposure is associated with immunosuppression, which may reduce the vaccine antibody response and increase the risk of infection. Studies have found that the level of PFASs is positively correlated with inflammatory markers such as C-reactive protein (CRP) in the serum, indicating that they may promote inflammatory responses. Regarding the relationship between CRP levels and the permeability of the blood–brain barrier, current research mainly indicates a statistical correlation rather than a strict causal relationship. Atmospheric exposure to PFASs may indirectly lead to an increase in CRP by inducing an inflammatory response and may affect blood–brain barrier function, but this mechanism still lacks direct support from animal experiments and longitudinal population studies. Future research needs to combine experimental verification and epidemiological data to further clarify the role played by atmospheric PFAS exposure in inflammatory responses and neurotoxicity risks. Chronic inflammatory states are closely related to various endocrine diseases, such as hypothyroidism, diabetes, polycystic ovary syndrome, etc., and chronic inflammation or immune activation induced by PFASs may be an important way in which they indirectly interfere with hormone secretion [56]. Experiments have also shown that PFASs can cause thymus atrophy and a decrease in lymphocyte levels, affecting immune function [57], particularly related to the liver and kidneys, which are important organs in the immune system. It has been found in the literature that the concentration of PFOS is relatively high in the liver (with a median of 41.9 ng·g⁻¹), while that of PFOA is relatively high in the bones (with a median of 20.9 ng·g⁻¹). These accumulations may potentially interfere with the immune system.

Exposure of the human body to atmospheric PFASs can cause toxicity in the nervous system. Exposure to PFASs during the developmental period may lead to neurobehavioral abnormalities, such as a decline in learning and memory ability and spontaneous behavioral changes. In the serum, PFOA (7.4 ng·mL⁻¹), PFOS (6.8 ng·mL⁻¹), and 6:2 Cl-PFESA (6.2 ng·mL⁻¹) are the main PFASs present, accounting for 79% of the total PFAS burden. Furthermore, in the CSF, PFOA (0.078 ng·mL⁻¹), PFOS (0.028 ng·mL⁻¹), and 6:2 Cl-PFESA (0.051 ng·mL⁻¹) are the main PFASs observed, accounting for 73% of the total PFAS burden. PFASs can penetrate the blood–brain barrier, accumulate in brain tissue, and interfere with the development of neurons [56]. The permeability of PFASs (RPFAS) was positively correlated with the brain barrier permeability index RAlb (CSF/serum albumin ratio) (r² > 0.6, p < 0.001), indicating that the integrity of the brain barrier is the main determinant of PFAS permeability. The level of CRP in the serum was positively correlated with the RPFAS value of the main PFASs, indicating that inflammation may cause PFASs to cross the brain barrier. For example, for PFOA, for every additional logarithmic unit of CRP, the RPFAS increases by 0.109 (95% CI: 0.036, 0.182).

PFASs in the atmosphere can also be toxic for reproduction and development. PFASs may affect reproductive health, such as reducing sperm quality and interfering with sex hormone levels [52]. Exposure among pregnant women may lead to fetal growth restriction and decreased birth weight [57]. This research also found that the accumulation of PFASs is related to age, and the concentration of PFASs in people over 60 years old is generally higher than that in other age groups. Smoking habits have a certain impact on the accumulation of PFASs in the lungs. The accumulation of PFASs in the lungs of smokers is lower than that of non-smokers. Some PFASs, such as 6:2 Cl-PFESA, have a stronger activating effect on the PPAR signaling pathway than PFOS and have higher potential developmental toxicity [56]. In the liver and brain, the concentration of PFHxA is the highest (with a median of 68.3 ng·g⁻¹ in the liver and 141 ng·g⁻¹ in the brain), while in the bones, PFOA is the main contributing substance (with a median of 20.9 ng·g⁻¹). The concentration of accumulated PFASs is the highest in lung tissue, but PFOS and PFOA are more common in the liver and bones.

The assessment of the human health risks posed by PFASs in the atmosphere requires comprehensive consideration of factors such as exposure levels, receptors, toxicological characteristics, and dose–response relationships. The health risks posed by PFASs in the atmosphere to the human body are not a single additive effect, but rather a synergistic effect of different compounds, meaning they may be much higher than the estimated value. For the human body, air and dust intake through breathing is the main source of PFAS. The action processes of the two routes, skin contact exposure and respiratory intake, involve many parameters, making it difficult to accurately measure the level of PFAS intake and exposure. It is worth noting that there are currently no specific exposure concentration limits for atmospheric PFASs, but there are toxicological parameters that can provide a preliminary reference for assessment. The U.S. Environmental Protection Agency (EPA) has set oral reference doses (RFDS) for perfluorooctanoic acid (PFOA) and perfluorooctane sulfonate (PFOS) at 2.0 × 10⁻⁵ and 2.0 × 10⁻⁵ mg/kg·day, respectively (based on the 2016 Health Recommendations) [50]. The Poison and Disease Registry (ATSDR) [49] of the United States has proposed a more conservative minimum risk level (MRL) for PFOA, PFOS, perfluorohexane sulfonic acid (PFHxS), and perfluorononanoic acid (PFNA) ranging between 2 × 10⁻⁶ and 3 × 10⁻⁶ mg/kg·day. Suppose an adult (70 kg) inhales 20 m³ of air daily. Under the condition of complete absorption, calculated based on the ATSDR MRL of PFOA (3 × 10⁻⁶ mg/kg·day), the corresponding atmospheric concentration is approximately 10,500 pg/m³. The observational data summarized in this review show that the concentration of PFASs in the air in most areas is far below this range. However, in some areas near pollution sources (such as the industrial area of New Jersey), the observed concentration may approach or exceed these reference levels, indicating potential health risks. As there is currently a lack of systematic efforts to set atmospheric exposure limits, future research and risk management should focus on this aspect.

4.2. The Impact of PFASs in the Atmosphere on Ecology

4.2.1. The Toxic Mechanism of PFASs in the Atmosphere on Plants

PFASs in the atmosphere can affect photosynthesis by influencing the integrity of plant cell membranes. For instance, PFASs can interact with phospholipids in plant cell membranes, altering the fluidity and permeability of these membranes, thereby influencing the synthesis of photosynthetic pigments and the efficiency of photosynthetic electron transport chains [58]. PFASs may also damage the structure and function of photosynthetic organs by inducing oxidative stress responses within plant cells. For instance, PFOS (perfluorooctane sulfonate) can increase the reactive oxygen species level within plant cells, leading to damage to the photosynthetic membrane and the inactivation of photosystem II [59].

PFASs in the atmosphere can also affect plants’ nutrient absorption and transport by binding to receptors on the surface of plant cells. For instance, PFASs can compete for binding sites with ion channels on these cells, thereby inhibiting the absorption of nutrients such as calcium and magnesium. PFASs may also affect the absorption of water and nutrients by plants by altering the morphology and function of their root systems. For instance, PFASs can inhibit the growth of plant roots, resulting in a reduction in the surface area of the roots and thereby lowering the efficiency of plants in absorbing nutrients from the soil [60].

4.2.2. Hazards to Animal Health

The concentration level of PFASs in the atmosphere is closely related to their cumulative effect within organisms. Birds can be indirectly exposed to PFASs in the air through respiration and the food chain. Precursors (such as FTOHs and FOSAs) can be converted into more persistent PFOS in the body, leading to its accumulation. Previous studies have shown that the LOAEL of PFOS in birds is approximately 50–100 ng·g⁻¹ wet weight [61]. In field investigations, air deposition and insect feeding are considered key sources. Similarly, the exposure of marine mammals is mainly achieved through deposition at the atmosphere–ocean interface and food chain transfer. For instance, in the Arctic region [62], despite the lack of direct emission sources, PFAS concentrations of tens to hundreds of ng·g⁻¹ have been detected in polar bears and seals [63], which is closely related to the cumulative effect of long-distance atmospheric transport. Therefore, linking the concentration levels in the air with data for within organisms can help to reveal the health hazards posed by PFASs to animals.

Exposure to PFASs may lead to impaired immune system function in animals, with studies showing that exposure to PFOS can affect the immune system of birds, causing liver and spleen enlargement and influencing the development of the wings and brain of embryos [64]. PFASs may also have a negative impact on the reproductive system of animals, with PFOS exposure reducing the hatching rate of birds, especially under exposure to high concentrations [65]. In addition, exposure to PFASs may affect the normal development of embryos and larvae, with studies finding that exposure to PFOA increases the incidence of leg deformities in poultry embryos and inhibits the development of yellow feathers [66]. Finally, exposure to PFASs may lead to liver damage, manifested as liver enlargement and abnormal liver function, as studies have shown that exposure to PFOS and PFOA can cause pathological changes in the livers of birds [67].

The assessment of PFASs’ impact is based on the toxicity reference value (TRV) and the minimum observed effect level (LOAEL). Studies have shown that the minimum observed effect level (LOAEL) of PFOS in birds is 100 ng·g⁻¹ [68]. The framework for the use of the TRV is explicitly stated in the risk assessment guidelines of Health Canada [69] and USGS [70] and has been applied in the recent EPA Human Health Toxicity Assessment document [71,72]. A concentration higher than this may lead to a decrease in the hatching rate [65]. The toxicity reference value (TRV) of PFOS is 1.7 μg·mL⁻¹, and the predicted no-effect concentration (PNEC) is 1 μg·mL⁻¹ [73]. The toxicity reference value of PFOA is relatively low, but its effect on embryonic development is more significant at high concentrations [66].

4.2.3. The Pathways Through Which PFASs in the Atmosphere Affect Ecological Risks

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Widespread distribution in the environment: PFASs are ubiquitous in the air, dust, water bodies, and living organisms, and indoor environments (such as carpets and textiles) are significant sources of pollution [34]. Firstly, for long-distance transmission and accumulation, PFASs are transported through the atmosphere and ocean to remote areas such as the Arctic and islands in the Indian Ocean. These substances accumulate in the marine ecosystem and may have long-term effects on marine life [74]. Secondly, studies have shown that plastic pollution can serve as an important carrier for PFASs, which can adhere to plastic particles and spread to remote ecosystems through marine plastic pollution [75].

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Food chain transmission: PFASs accumulate through aquatic and terrestrial food chains, reaching high levels in predators at high trophic levels, thus affecting the survival and reproduction of wild animals such as fish and birds [57]. Studies have shown that PFASs have a significant biological amplification effect in the marine food chain, especially in seabirds and marine mammals [73], and accumulate in the marine ecosystem, potentially having long-term effects on marine life [74]. Studies have also shown that the concentration of PFASs in seabird eggs may affect the development and hatching success rate of embryos [76].

5. Sample Collection Methods and Sampling Techniques of PFASs in the Atmospheric Environment

The sampling methods for PFASs in the air mainly include active air sampling methods and passive sampling methods.

Active air sampling methods involve using an air extraction device to pass air through a filter membrane, effectively collecting PFASs attached to particulate matter. These methods can simultaneously collect PFASs in both the gas phase and the particulate phase and have a relatively high sampling efficiency. The gas-phase PFASs are enriched onto polyethylene and other films through a vacuum pump and then captured by adsorbents such as PUF or XAD resin (where PUF stands for polyurethane foam and XAD is a non-polar macroporous resin). This method has the advantages of controllable flow and high sampling efficiency and is suitable for short-term and quantitative research. During the sampling process, a high-flow sampling pump, a glass fiber filter membrane, and an adsorption cylinder should be used, and the flow velocity should be calibrated in combination with the flowmeter. After collection, the sample needs to be refrigerated and stored and quantitatively detected through solvent extraction and LC-MS/GC-MS analysis (Figure 4).

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Gaseous PFASs: Collected using polyurethane foam (PUF) or activated carbon adsorption tubes. Avoid using sampling equipment containing fluorine materials.

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Particulate matter: PM_2.5/PM₁₀ is collected through a quartz-fiber filter membrane and then analyzed by solvent extraction.

The passive sampling method [77] has shown great potential in monitoring pollutants. By integrating sampling and analysis, the sensitivity is greatly improved, and through continuous deployment for several days to several weeks, the time-weighted average (TWA) concentration of pollutants over a period of time can be obtained. These methods can thus be used for the detection of trace or ultra-trace pollutants in the environment.

Passive samplers are divided into two types: diffusion samplers and penetration samplers. Adsorbents come in various types, mainly determined by the substance to be tested. Common passive sampling techniques include semi-permeable membrane devices, chemical traps, integrated samplers for polar organic substances, ceramic dosimeters, and solid-phase microextraction and thin-film diffusion gradient technology.

The analysis and detection of PFASs mainly rely on liquid chromatography–tandem mass spectrometry (LC-MS) and gas chromatography–mass spectrometry (GC-MS) techniques (

Table 4. The principle and application of the analytical methods for the detection technology of PFASs.

Detection Method	Principle	Application	Reference
LC-MS	Separation by liquid chromatography, detection in ESI− mode	Broad PFASs detection (water, soil, air)	[78]
GC-MS	Volatile/semi-volatile PFASs analyzed after derivatization	FTOHs, FOSAs, volatile PFASs in air/textiles
TOPA	Oxidation of precursors → stable end products	Indirect quantification of PFAS precursors
TOF	Measure total fluorine content	Screening for unknown PFASs

). LC-MS separates PFASs through liquid chromatography and adopts the negative-ion mode of the electrospray ion source for tandem mass spectrometry analysis. For liquid chromatography (LC), when the sample passes through the chromatographic column, different components are separated due to their varying interactions with the stationary phase. In mass spectrometry (MS), the separated components enter the mass spectrometer, are ionized, and are detected and quantified according to the mass-to-charge ratio (m/z). Common ionization methods include electrospray ionization (ESI) and atmospheric pressure chemical ionization (APCI). LC-MS displays high sensitivity and selectivity and is widely used in the detection of various PFASs in water, soil, and biological samples. Separation is carried out by taking advantage of the difference in the distribution coefficients of solutes between the stationary phase and the mobile phase. When a sample solution containing multiple components enters the chromatographic column, due to the differences in adsorption, dissolution, distribution, and other characteristics of each component between the stationary phase and the mobile phase, their movement speeds in the chromatographic column are different, thus achieving the separation of each component.

GC-MS is suitable for the analysis of volatile or semi-volatile PFASs such as fluorinated alkanols and fluorinated sulfonamides. The samples need to undergo derivatization treatment before gas phase separation and then are detected by means of mass spectrometry. This technique is often used for the analysis of PFASs in the atmosphere. In gas chromatography (GC), after the sample is vaporized, it is carried into the chromatographic column by the carrier gas and separated based on the differences in boiling point and polarity. In mass spectrometry (MS), the separated components are ionized by electron bombardment (EI) or chemical ionization (CI) to generate characteristic ion fragments. Volatile or semi-volatile compounds (such as FTOH, FOSA, etc.) are separated by GC and detected by MS. GC-MS can directly detect PFASs in indoor and outdoor air, while also being able to analyze and detect volatile PFASs in paper and textiles. However, GC-MS is not suitable for ionic or thermally unstable PFASs, as the sensitivity of this technique is usually lower than that of LC-MS. GC-MS is therefore a supplementary technology specifically designed for the analysis of volatile PFASs.

Other auxiliary technologies include TOPA and TOF. TOPA is the process of converting unknown PFAS precursors into detectable final products using strong oxidants such as persulfates. TOF is used to determine the total fluorine content by means of combustion ion chromatography (CIC) or particle-induced gamma-ray emission (PIGE). Both of these methods enable the quantitative analysis of PFAS precursors in environmental samples, filling the gap in the reference standards.

6. Pollution Control and Treatment of PFASs in the Atmosphere

6.1. The Particularity of PFAS Control in the Atmosphere

The treatment of PFASs in the atmosphere has significant particularities and challenges, mainly due to their unique physicochemical properties, environmental behaviors, and technical limitations. The text above also described why PFASs have chemical stability and persistence, and their long-distance migration and global distribution. The diverse emission sources and complex existence forms of PFASs are important reasons for its particularity. It is not only the dispersion of sources: industrial emissions, consumer product volatilization, waste incineration, etc., that makes monitoring and control difficult, but PFASs also coexist in gaseous, particle-adsorbed, and water-soluble ionic forms, and require the coordinated treatment of multiple technologies. The text above also mentioned that precursors in the atmosphere can generate more stable end products through photolysis or oxidation, which further increases the complexity of governance. Some of the pollution control technologies introduced below have limited efficiency for nanoscale PFASs in the atmosphere. Traditional dust removal or adsorption technologies need to be refined into advanced methods such as efficient catalytic degradation or low-temperature plasma. Finally, it is not difficult to see that the pollution control and treatment strategies for atmospheric PFASs have the challenge of cross-border supervision, and international collaboration is still needed to manage the emission sources, especially for industrial emissions and the use of consumer products containing PFASs. Therefore, the governance of atmospheric PFASs requires a full-chain strategy of “source–process–end”, emphasizing the combination of forward-looking monitoring and innovative technologies.

6.2. Common Methods for Controlling Atmospheric PFASs Pollution

Although controlling atmospheric PFASs pollution is difficult, the following technologies show potential breakthroughs.

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Physical methods (adsorption, filtration) [79]

Physical methods include activated carbon adsorption, ion exchange resins, and new types of adsorption, etc. Activated carbon (AC) is the most widely studied adsorbent and has a good adsorption effect on long-chain PFASs (PFOS and PFOA). Both granular activated carbon (GAC) and powdered activated carbon (PAC) can effectively remove PFASs, but they need to be replaced or regenerated frequently. Ion exchange resins (such as the Amberlite series) have a relatively high removal efficiency for PFASs, especially for negatively charged PFASs (PFOS and PFOA). The resin can be regenerated through salt solutions such as NaCl and methanol. Other adsorbents include molecularly imprinted polymers (MIPs), carbon nanotubes (MWCNTs), magnetic mesoporous carbon–nitrogen materials (MMCN), etc. Some of these materials exhibit highly efficient adsorption capabilities under specific conditions. This method has a relatively high removal efficiency for particulate PFASs. For instance, when treating long-chain PFASs (such as PFOS) with anion exchange resin (IEX), the removal rate can be as high as or exceed 92%, while the effect on short-chain PFASs is relatively poor [80]. This method is particularly efficient in removing granular contamination, but its ability to act on gas-phase PFASs is significantly insufficient.

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Chemical methods (advanced oxidation, thermal decomposition) [81]

The main chemical methods are photocatalysis and thermal decomposition. Photolysis and photocatalysis utilize ultraviolet irradiation to degrade PFASs through direct photolysis or photochemical oxidation. Direct photolysis has a poor effect on PFASs and requires specific wavelengths (such as VUV < 190 nm), while the combination of chemical reagents/catalysts and ultraviolet light can effectively degrade it. For example, the UV–Fenton process has a certain removal and defluorination effect on PFOA. However, this method has a poor effect on PFOS, and the degradation and defluorination efficiency is relatively low. The C-F bond of PFASs in the atmosphere is more difficult to be completely broken. Low-temperature plasma (NTP) can generate highly active free radicals (·OH, O₃) to decompose PFASs, but it has relatively high energy consumption. Thermal decomposition is essentially the incineration of PFASs: high temperatures (600–1000 °C) can completely decompose PFASs, but it may produce harmful by-products (such as hydrogen fluoride and greenhouse gases), and it also requires the installation of corresponding tail-gas treatment equipment, which is costly. Chemical methods have shown high degradation efficiency for both gas-phase and granular PFASs. For instance, the treatment of PFOA by UV–Fenton showed that the removal rate reached 87.9% within 1 h, and the total removal rate exceeded 95% after 5 h, with the fluoride removal efficiency rising to 53.2% [82]. Another one employs a UV/S + I system, achieving a removal rate of over 99.7% for various PFSAs and PFCAs, with a fluoride removal efficiency exceeding 90% [83]. However, these processing procedures may generate by-products such as short-chain PFASs or fluorine-containing intermediates. Non-thermal plasma technology has recently been used in experiments to degrade PFAS substitutes such as ADONA and GenX, and the degradation pathways and by-products have been preliminarily described. However, the toxicity of the final products remains unknown [84].

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Biological methods (microbial degradation) [85]

Microbial degradation of PFASs usually begins with the enzymatic attack of the C-F bond by reductase for reductive defluorination, where fluorine atoms are replaced by hydrogen to generate intermediate products such as some fluorinated hydrocarbons. However, due to the fluorine shielding around the carbon–carbon bonds in PFASs molecules, this process is hindered, making it difficult for microorganisms to utilize PFASs as a carbon source and energy. Many microbial strains have demonstrated certain capabilities in the degradation of PFASs. For example, Acidimicrobium sp. A6 can degrade PFOA and PFOS by 60% within 100 days under anaerobic conditions, and produce transformation products such as PFBA and PFPeA. In another study [86], this strain enabled PFOA to degrade by 50% within 150 days. Under aerobic conditions, Pseudomonas aeruginosa strain HJ4 can reduce the concentration of PFOS by 67% within 48 days, and the transformation products are PFBS and PFHxS. Gordonia sp. Strain NB4-1Y can degrade 99.9% of 6:2 FTSA within 7 days and generate various transformation products. These data indicate that different microorganisms have varying degradation capabilities and transformation products for different PFASs. At present, this observation is mainly based on laboratory conditions, and although microbial transformation has some degradation capacity, its efficiency and practicality are significantly limited. A review points out that microbial degradation of PFAS is often accompanied by long reaction times, extremely low degradation rates and complex unknown by-products. To date, there is no completely defluorinated microbial system that can be applied in the real environment [87].

7. Conclusions and Prospects

7.1. Conclusions

Through a review and analysis of the classification and properties, pollution characteristics, and migration processes of PFASs in the atmosphere, along with related toxicological analysis and assessment, detection techniques, and pollution control measures, the following conclusions are obtained.

7.1.1. Key Points of the Current Research Status of PFASs

PFASs are known as “permanent chemicals” due to the presence of strong C-F bonds and exhibit significant persistence and mobility in the atmosphere. Short-chain PFASs are more likely to exist in a gaseous form and be transmitted over long distances, while long-chain PFASs are more often adsorbed onto particulate matter and undergo stronger bioaccumulation. Monitoring results show that there are significant differences in atmospheric PFAS concentrations, both globally and among regions in China. For instance, the particulate phase ∑PFAS concentrations in Beijing and Chengdu reached 157 pg·m⁻³ and 140.81 pg·m⁻³, respectively, with short-chain PFBA accounting for as much as 69% of the total burden in Chengdu. The proportion of FTOHs in the South China Sea gas phase exceeds 95%, indicating that the input of continental air currents contributes significantly to this. In contrast, the concentration of gas-phase PFASs (197.7 ± 47.9 pg·m⁻³) in New Jersey, USA, exceeded that of the granular phase (48.3 ± 47.9 pg·m⁻³), indicating that short-chain substitutes have become a major source of risk in North America. Studies conducted in the polar regions and Northern Europe have also found that the gaseous phase accounts for as much as 91% of the overall burden, reflecting the long-distance migration ability of PFASs. Overall, the southern coastal areas of China are mainly affected by short-chain gaseous substances, while the concentration of particulate substances is higher in the north and around industrial cities. In some cities (such as Zhejiang, 251.93 pg·m⁻³), the concentration has approached or exceeded that of developed countries in Europe and America.

Toxicological evidence indicates that atmospheric PFASs pose significant risks to humans, animals, and plants. In humans, PFOA is associated with an increased risk of liver cancer, testicular cancer, and kidney cancer. Animal experiments have shown that a PFOS inhalation LC₅₀ of 5.2 mg·L⁻¹ and a PFOA inhalation concentration of 980 ng·m⁻³ can cause respiratory and liver damage. Epidemiological studies have also found that PFOA and PFOS can cross the blood–brain barrier, accumulating to 2 ng·g⁻¹ and 5.3 ng·g⁻¹, respectively, in thyroid tissue, and are associated with glucose metabolism disorders (β ≈ 0.320–0.392, p < 0.05) and immunosuppression. In terms of ecology, the minimum observed effect level (LOAEL) of PFOS on birds is 100 ng·g⁻¹. Exceeding this value will significantly reduce the hatching rate. Plants, on the other hand, show impaired photosynthesis and inhibited root development in response to PFASs. Overall, atmospheric PFASs pose a lasting threat to humans and ecosystems through multiple pathways and multi-organ effects.

Although detection methods such as LC-MS/GC-MS and treatment methods such as physical adsorption, chemical oxidation, and microbial degradation have been continuously advanced, the existing technologies are limited by their sensitivity, efficiency, and cost, meaning that the efficient control of atmospheric PFASs is difficult to achieve.

7.1.2. Regional Perspectives (Urban/Coastal/Industrial/Polar)

Summary of regional characteristics: As mentioned above, existing research has revealed the distribution differences of atmospheric PFASs in different regions: in urban areas, short-chain PFASs and indoor and outdoor exposure are prominent issues. Coastal areas are significantly affected by the exchange between the atmosphere and the ocean interface. Point-source emissions from industrial zones lead to local high-concentration accumulation. The monitoring results in polar and remote areas highlight its long-distance migration capability. These differences provide important regional references for subsequent research and governance.

7.2. Prospects for Future Research

At present, scholars around the world are breaking through traditional cognitive limitations through multi-media coupling models and ultra-high-resolution mass spectrometry technology. Furthermore, policymakers are highly concerned about the challenges that cross-media circulation poses to the management of “permanent chemicals”. Current research on PFASs in the atmosphere focuses on four core directions: (1) the construction of long-distance atmospheric migration mechanisms and global diffusion models; (2) assessing the environmental behavior and secondary pollution risks of new substitutes; (3) establishing the dynamics of gas/particle distribution and the influence of interfacial reactions on the fate of PFASs; and (4) the development of traceability technology based on isotope fingerprints and regional contribution analysis. It is suggested that future research focus on the following directions:

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Developing predictive models incorporating multi-phase interactions and deposition processes. Future models should move beyond simple transport models to integrate dynamic gas/particle partitioning, interfacial chemical reactions, and dry/wet deposition processes. This will enable the more accurate prediction of the environmental fate of PFASs, from emission sources to ultimate sinks in remote areas, and support the development of targeted mitigation strategies.

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Proposing targeted mitigation methods for specific sources. Research should focus on developing efficient and economically feasible control technologies for major emission sources. This includes (1) end-of-pipe treatment technologies for industrial exhaust gases, such as non-thermal plasma coupled with catalysts or customized adsorbents for volatile precursors; (2) source control strategies, such as finding alternatives to PFASs in consumer products that are prone to volatile release; and (3) exploring remediation techniques for secondary sources, such as suppressing the re-release of PFASs from contaminated soils into the atmosphere.

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Strengthening long-term atmospheric monitoring systems and building a comprehensive monitoring network covering typical regions (industrial, urban, coastal, remote). Research should utilize both active and passive sampling techniques to obtain large-scale, long-term temporal and spatial data, clarifying regional pollution trends and providing a robust data foundation for model validation and policy formulation.

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Improving exposure and health risk assessment frameworks, integrating atmospheric concentration data with multi-pathway exposure models (inhalation, dust ingestion, dermal contact) to establish a composite risk assessment system, and strengthening epidemiological and toxicological mechanistic studies to clarify the dose–response relationship and pathogenic mechanisms of low-concentration, long-term exposure to atmospheric PFASs.

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Enhancing institutional and global collaboration, promoting the improvement of policies and regulations covering the entire lifecycle of PFASs, reducing transboundary pollution through global agreements, and strengthening international cooperation in scientific research, data sharing, and risk management to jointly address the challenges posed by these persistent pollutants.

Regionalization Outlook:

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Urban areas: Strengthen long-term air monitoring and residents’ exposure risk assessment, and explore the inclusion of PFASs in the existing air quality regulatory framework.

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Coastal areas: Focus on the material exchange and deposition fluxes at the atmosphere–ocean interface, and evaluate their input effects on marine ecosystems in combination with models.

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Industrial zone: Improve source apportioning and emission inventories, and focus on developing control and alternative assessment technologies suitable for industrial waste gas.

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Polar and remote regions: Establish long-term observation stations, combine models to verify the long-distance atmospheric transmission mechanism, and reveal its accumulation and risks in the original ecosystem.