Occurrence of Microplastics in the Atmosphere: An Overview on Sources, Analytical Challenges, and Human Health Effects

Abstract

The rapid spread and accumulation of microplastics (MPs) in environmental ecosystems result from extensive plastic usage. MPs have been found in both indoor and outdoor air. Outdoor MP levels vary widely across global cities, with reported ranges from 36 to 118 MPs m⁻² day⁻¹. However, differing measurement units complicate comparisons. Indoor MPs are particularly concerning due to the significant amount of time people spend indoors. For instance, MP concentrations in workplaces like reception areas and nail salons were found to be 309 ± 214 and 46 ± 55 MPs m⁻³, respectively. Technological limitations hinder the identification of MPs, with methods like µ-ATR-FTIR, µ-FTIR, and µ-Raman identifying MPs of different sizes. MPs smaller than 0.3 µm pose a health risk as they can be internalized in lung cells, while MPs larger than 10 µm are too large to enter alveolar macrophages. This review highlights the current understanding of airborne MPs, focusing on their sources, transport, and deposition mechanisms. It aims to provide a foundation for further studies to deeply assess the presence, abundance, and occurrence of MPs in aerosols, a subject that remains underexplored.

1. Introduction

Plastic, widely used in various products globally, has become a significant environmental concern due to its extensive use and poor management practices [1,2]. The repercussions are evident in the rapid accumulation of approximately 8 million tons of plastic in the marine ecosystem annually [3]. Here, plastic materials are subject to biotic and abiotic degradation, processes like photooxidation and abrasion. Such mechanisms could lead to the loss of their mechanical integrity. These processes result in the emissions of microplastics (i.e., MPs) [4,5]. Currently, MPs are a relevant class of anthropogenic contaminants. MPs have become a global concern for research all around the world, mainly due to their potential negative and adverse effects on organisms (humans) and ecosystems [6,7]. Depending on their formation process, MPs are clustered into two major groups: primary MPs, which are directly emitted from the production processes of some products (i.e., scrubs), and secondary MPs, which originate from degradation or fragmentation of macro-sized plastic litter or debris. In both cases, the upper dimensional limit is 5 mm [8,9] as defined by Thomson et al. in 2004 [10]. Thus far, MPs have been detected across various environmental matrices, including marine vertebrates and invertebrates [11], terrestrial organisms [12], as well as human tissue [4]. The atmosphere has gained particular attention, as it has been identified as one of the main source of MPs exposure for organisms [13,14]. Due to their small size and low density, airborne MPs can be transported by winds and transferred between aquatic and terrestrial environments, spreading globally [2,15,16,17,18,19]. Airborne MPs originate from several sources, including personal care products, clothing, household items, packaging, paints, 3D printers, and road dust [4,20]. In urban settings, dust and road traffic, through processes such as tire and brake rubber abrasion, paint particles from road markings, asphalt, and road markings, contribute significantly to airborne MPs [20,21,22].

Based on the remarkable occurrence of MPs in the air, effects of airborne MPs on human health have been studied. MPs can be, in fact, easily inhaled by humans, and thus pose a potential threat on human health [23]. Concerns arise regarding the potential health effects of inhaling airborne MPs as studies show their presence in human lung tissues (particles size ranged from 1.60 to 5.56 μm and fiber size ranged from 8.12 to 16.8 μm) [4], confirming that inhalation plays an important role in human exposure to MPs [2,4,14,19]. However, studies have only confirmed the presence of MPs and not their absorption by the cells of the lung tissues. Moreover, their toxicological effects remain still unclear, as up to now, there is not enough evidence to state that airborne MPs are toxic to human health via inhalation [1,2,14,19]. Uncertainties about MPs’ effects on human health are mainly related to the lack of analytical methodologies. Analytical methodologies for the quantification and qualification of MPs are crucial to understanding their source of pollution and assessing environmental risk [24]. To assess the impact of airborne MPs accumulation in human tissues in depth, concentration, shape, and, mostly, the particle sizes of MPs in the atmosphere need to be clearly assessed [7,18,25,26]. It has been shown that fine particles, defined as particles with an aerodynamic diameter smaller than 2.5 µm (PM2.5), can enter the respiratory system and cause health damage [27,28]. More precisely, PM_0.3 is the most penetrating particle size and is highly harmful due its significant capacity to carry bacteria and viruses [29]. For this reason, due to their size, MPs with a diameter smaller than 2.5 µm could also enter the respiratory system. Negative and long-term effects of MPs, in fact, require their accumulation in tissues. Accumulation of MPs into lung tissues requires the embedding into cells’ systems [14]. It has been reported that only MPs smaller than 2.5 µm could be adsorbed into tissues through the respiratory system [4,30]. Therefore, studies that deal with accumulation of MPs through the respiratory system and consequent negative effects should consider a particle size range with biological plausibility.

The present review aims to provide a comprehensive overview of current knowledge on airborne MPs, emphasizing sources, transport, and deposition mechanisms. Furthermore, the possible impact on human health will be discussed. This study will also focus on current analytical methods, identifying analytical issues for sampling and the analysis of samples. This can provide a base for further studies towards in-depth assessment of the presence, abundance, and occurrence of MPs in the aerosol, which are still under investigation.

2. Materials and Methods

For the purpose of this review, search terms were used for each concept for relevant articles. The search terms for each concept were combined using the Boolean operator “OR”, while the categories (concepts) were combined using the Boolean operator “AND” (

Table 1. Search terms used for the literature search.

Concept 1	Concept 2	Concept 3
Atmospheric microplastic *	Analytical determination *	Human health
Atmospheric plastic particle *	Sample treatment *	Adverse effect *
Atmospheric microplastic particle *	Analytical method *	Inhalation
OR	OR	OR
AND		AND (Mesh)

* Indicates that both singulars and plurals were considered for the research.

).

A total of 137 articles were identified through the Scopus and Google Scholar databases, using specific search terms reported above.

Articles that were obtained went through a two-step screening process based on the inclusion (Table 1) and exclusion criteria. The first step involved evaluating the title and abstract of each publication, and the second step involved examining the full text of each article. As a result, only articles that met the inclusion criteria have been included in this review. Among the papers retrieved, studies on the presence of MPs in dust or sediments and studies on exposure through ingestion were not included in this review.

Out of these, 37 articles were excluded based on the screening of their titles and abstracts. After reading the full text of the remaining 100 papers, 80 were found to be relevant and included in this review. This review included articles that examined the assessment of microplastics in both indoor and outdoor environments, discussing the analytical challenges and the potential impact on human health. Articles included provided information about the occurrence of microplastics in the atmosphere, as well as the possible sources of contamination. Moreover, this review also considered articles that offered valuable insights into the identification of shortcomings in the analytical methods used. Articles considered potentially relevant were thoroughly read if their titles and abstracts indicated their relevance. Detailed information regarding the occurrence and abundance of microplastics in the atmosphere, potential sources of contamination, analytical methods used for pre-treatment and quantification, and the impact of microplastics on human health were gathered whenever they were available.

3. Results

3.1. MPs in the Atmosphere

The scientific community has only recently provided attention to airborne MPs [31]. For instance, research by Dris et al. (2015) was the first to suggest the existence of MPs in the atmosphere in Paris (France). In particular, it was reported that an average of 118 particles m⁻² day⁻¹ (90% of them were classified as fibers) were present in the outdoor atmosphere [32]. Similarly, Cai et al. (2017) characterized MPs in the atmospheric fallout in Dongguan city (China). Polyethylene (PE), polystyrene (PS), and polypropylene (PP) MPs at an average level of 36 ± 7 particles m⁻² day⁻¹ were detected. No information about size range was reported [33] (

Table 2. Summary of data reported by included studies where atmospheric MPs have been identified, quantified, and characterized in remote, urban, and suburban areas. Abbreviations: polyethylene terephthalate (PET), polyvinyl chloride (PVC), polytetrafluoroethylene (PTFE), nylon (NY), polymide (PI), polyacrylate (PA), polyurethane (PUR), rayon (RY), polyethersulfone (PES), polymethyl methacrylate (PMMA), polyoxymethylene (POM), acrylonitrile (AN), epoxy (EP), polyacrylonitrile (PAN), poly vinyl acetate (PVA).

Location	Area	Sampling Periods	Average Concentration	Shape	Polymer Type	Ref.
Asaluyeh, Iran	City country	August 2017	1 MP m−3	Fiber	N/A	[38]
Shiran, Iran Mount Derak	Urban area Remote area	1 October 2019 to 30 September 2020	1909 ± 1038 MPs m−2 366 ± 142 MPs m−2	Fiber	PP, PS, PE	[31]
Shiraz, Iran	Urban area	28 December 2021 to 5 January 2022	20 MP m−2 h−1	Fiber Fragment	PET, PVC, PTFE, PS, PP, PE, NY	[39]
Bushehr port, Iran	Urban area	December 2016 and September 2017	2.1 items m−3 (normal days) 10.3 items m−3 (dusty days)	Fiber, Fragment, Film	PET, PE, NY, PS	[6]
French Pyrenees	Remote area	November to March	365 ± 69 m−2 day−1	Fragment, Fiber, Film	PS, PE, PP	[34]
Sao Paulo, Brazil	Urban area	24 September to 1 November 2020	123.20 ± 47.09 MPs m−2 day−1	Fragment, Film, Granule, Foam	PL, PE, PET	[40]
Ankara, Turkey	Urban areas	January to March 2022	59.66 items L−1	Fiber, Fragment, Film, Sphere	PE, PS, PP, PET, PVC, NY	[41]
Dongguan, China	Urban area	1 October to 31 December 2016	36 ± 7 MPs m−2 day−1	Fiber, Foam, Fragment, Film	PE, PP, PS, cellulose	[33]
Atlantic Ocean	Remote area	15 December 2020 to 14 January 2021	7.85 to 51.74 ng m− 3	Fiber	PE, PP, PI, PS	[42]
Gebze	Urban area	Each season *	4035 to 58,270 MP m−3	Fiber, Film, Fragment	PA, PE, PET, PP, PVC	[43]
Tempe, Arizona	Suburban area	28 October 2020 to 1 November 2021	0.2 MPs m−3	Fiber, Fragment	PVC, PL, PS, PE	[2]
Tainan and Kaohsiung, Taiwan	Suburban area	February to April 2021	28 ± 24 MPs m−3	Fragment, Fiber	acrylic, Rubber, PUR	[1]
Tehran, Iran	Urban area	5 July to 4 August 2019 and 8 November to 15 December 2019	0.8 N m−3	Fiber, Bead, Fragment	PP, PET, PS, RY, PVC, PE, PA	[44]
Northwestern Pacific Ocean	Remote area	26 September to 9 October 2017	0.027 ± 0.018 MPs m−3	N/A	RY, PET	[36]
Northwestern South China Sea	Remote area	24 October to 24 November 2019	0.035 ± 0.015 MPS m−3	Fiber, Fragment, Granule	PES, RY, PP, PE, PS, PA, PR	[35]
Paris, France	Urban area	Once in April and once in October 2018	118 particles m−2 day−1	Fiber	N/A	[31]
Paris, France	Urban area	19 February 2014 to 12 March 2015	110 ± 96 particles m−2 day−1	Fiber	RY, PET	[45]
	Suburban area	3 October 2014 to 12 March 2015	53 ± 38 particle m−2 day−1
Paris, France	Urban area	February, May, July, and October 2015	0.9 particles m−3.	Fiber	N/A	[46]
Baltic sea	Remote area	16 to 31 October 2019	171 ± 47 m−3	Fiber, fragment	PVC, PE, PET, PA	[37]
Scientific campus, Venice, Italy	Urban area	July 2021	N/A	N/A	PE, PS, PP	[47]
Krakow, Poland	Urban area	June 2019 to February 2020	N/A	Fiber, Fragment	PE, PP, PUR, PS, PET	[26]
Northeast Arabian Sea, India	Remote area	November 2020	1.46 ± 0.12 n m−3	Fiber, Fragment, Film	PVC, PMMA, PES, POM, PUR, AN	[48]
Weser River Catchment, Germany.	Remote area	April and October 2018	N/A	Spherical	PE	[30]
Hamburg, Germany	Urban area	December 2017 to February 2018	136.5–512 MP m−2 day−1	Fragment	PE	[49]
Christchurch, New Zealand	Suburban area	25 to 30 January 2020, 6 to 11 February 2020, 10 to 15 June 2020, 17 to 22 June 2020	80–330 particle m−2 day−1	Fiber, Films, Fragment	PL, PE	[18]
Beijing, China	Urban area	30 December 2020	123.6 items g−1	Fragment, Pellet, Fiber	PP, PA, PS, PE, PET, PVC	[8]
Lanzhou, China	Urban area	1 February to 31 August 2020	353.83 n m−2 d−1	Fragment, Fiber	PET	[50]
Harbin, China	Urban area	9 to 15 June 2021	1,76 n m−3	Fiber, Fragment, Granular	PP, PET, PE, PS, PVC	[15]
Beijing, China	Peri-urban farmland	1 September 2021 to 28 February 2022	167.08 ± 92.03 items m−2 d−1	Fiber, Fragment, Film, Foam	PET, RY, PP, PE, PA, PS	[51]
Red Sea, Saudi Arabia	Residential area	September 2015 to December 2017	0.9 ± 0.8 × 10−2 MFs m−3	Fiber	PET, PP, PE, NY	[52]
Brisbane, Australia	Rural location	27 April to 8 June 2020	0.53 ± 0.16 mg g−1	N/A	PP, PS, PET, PVC, PE	[22]
	Rural Residential		0.80 ± 0.49 mg g−1
	Residential		0.68 ± 0.20 mg g−1
	Industrial		2.4 ± 0.55 mg g−1
	Traffic		1.2 ± 0.70 mg g−1
	City		5.9 ± 3.1 mg g−1
Nanjing, China	Urban area	Monthly for 25–31 days	302.32 ± 107.40 items m−2 d−1	Fragment, Fiber	PA, PE, PES, EP, PP	[53]
Newfoundland, Canada	Remote area	9–11 September 2021	1.13 × 105 particles m−2 day−1	Fragment, Fiber	PMMA, PET, PVA, PES, PA	[54]
Pacific Ocean	Remote area	10 October 2019–5 January 2020	0.841 ± 0.698 items 100 m−3	Fiber, Fragment	PET, PA, PR, PP	[55]
London; United Kingdom	Urban area	19 January to 16 February 2018	575 to 1008 MP m−2 d−1	Fiber	PAN, PET, PA	[5]
Guangzhou, China	Urban area	January to December 2021	0.01 to 0.44 items m−3	Fragment, Fiber, Film, Pellet	PP, PET, PE, PA	[56]
Wuhan, China	Urban area	December 2020 to December 2021	82.85 ± 57.66 n m−2 d−1	Fiber, Fragment, Film, Pellet	PET, PP, PE, PA, PVC, PVA, PAN	[57]

* Thirty sampling filters were collected in each season. The period is not specified in the paper.

). However, the occurrence of microplastics was also confirmed in remote areas [31,34,35,36,37] (Table 2), suggesting their easy transport through various atmospheric phenomena. Table 2 summarizes all the data reported by the studies where outdoor atmospheric MPs have been studied in different areas.

Atmospheric MPs have also been investigated in indoor environments. The attention to indoor atmospheric MPs is based on the assumption that people spend most of their time in several indoor environments (i.e., offices, home). Research-based data seemed to reveal that the concentration of MPs transported in indoor environments through the air may be higher than that outdoors [13,20,45,49,58,59]. For instance, average fallout rates of 309.40 ± 214.71 MPs m⁻² day⁻¹ and 46 ± 55 MPs m⁻³ were detected in a reception room and in a nail salon, respectively [1,60]. An average of 123.20 ± 47.09 MPs m⁻² day⁻¹ and 28 ± 24 MPs m⁻³ were detected in the areas outside the reception room and nail salon, respectively, confirming the higher concentration of MPs in indoor environments than outdoors [1,60]. This can be due not only to the transport of MPs by air from outdoors to indoors, but also to synthetic fabrics, clothing, and household items that can release microplastics into the indoor air [20,40,58].

The amount of MPs in outdoor environments differs between areas. Geographical and temporal variations have been attributed to seasonal and climatic conditions, but also to variations in the primary sources of MP emissions and differences in sample collection and identification analysis [15,33,38,39].

3.1.1. Main Sources of Atmospheric MPs

Airborne MPs can be derived from industrial emissions, inadequate landfill management, incineration practices, dust resuspension, and abrasion of roads and tires. Additionally, MPs can also originate from fragments of clothing and household furniture [2,6,8,21,26,45]. PP, PE, PS, and PET are the primary polymeric components found in the atmosphere. These polymers are commonly used in different products details like plastic bags, packaging materials, textiles, furniture, and building materials [42,60,61,62]. Such plastics materials may be subject to UV degradation and physical abrasion, and fiber, fragment, film, or foam MPs may originate from them [5,22,38,63]. Several studies have shown that synthetic fibers torn from clothes and other fiber products during wearing, cleaning, and drying are the main sources of fibrous microplastics in air, both indoor and outdoor [30,31,32,33,43,45,64]. However, a relevant source of MPs, particularly in densely urbanized cities, is road traffic. Re-suspension of road dust and tire and road abrasion determine the release of MPs into air [21,22,31,65].

Indoor and outdoor MPs present some differences. The main difference observed is in their sizes. Dries et al. (2017) observed fibers ranging from 50 to 1650 µm in outdoor air and from 50 to 3250 µm in indoor air [46]. Similarly, Gaston et al. (2020) found that the sizes of fibers and fragments in outdoor air (fibers 25–2061 µm, fragments 51–408 µm) were smaller compared to those in indoor air (fibers 22–8961 µm, fragments 20–850 µm) [66]. It is possible that indoor MPs undergo less alteration (e.g., less exposure to UV rays, wind, temperature fluctuations, etc.) and are, therefore, somewhat less prone to fracturing compared to those in outdoor air. Moreover, the stable conditions and low ventilation inside buildings allow these larger fibers and fragments to settle more easily compared to the turbulent outdoor environment. Another difference is in the type of polymer. In indoor air, there is an abundance of PE and PET. These polymers are common in food packaging, construction materials, and textiles. Meanwhile, outdoors, the most abundant polymers are PVC and acrylic, often used in exterior coverings [66].

3.1.2. Fates of Airborne MPs

The presence of MPs in isolated regions, where anthropogenic activities are scarce, suggests their ability to be transported through the atmosphere. Similarly to other particulate matter, the atmospheric transport, concentration, and deposition of MPs are influenced by meteorological factors such as precipitation, wind speed, wind direction, pollutant density, and industrial activities. These factors can influence the levels of MPs in different environments, from urban areas to remote mountain and marine environments [1,5,6,8,15,21,26,30,56,60,61,64].

Regions with higher levels of human activity appear to have a greater concentration of MPs due to the numerous emission sources present. Once released, MPs can end up and remain suspended in the atmosphere, and they may be transported by the wind before being deposited through either wet or dry deposition [8,42,64,65,66,67]. In a transoceanic study, Wang et al. (2020) showed that MPs in the atmosphere can travel distances greater than 1000 km [59]. Ding et al. (2022) found that wind speed has a negative association with atmospheric MP concentrations, indicating that the deposition of MPs in the atmosphere is facilitated by low wind speeds due to their small size and low densities. The same study did not find a correlation between MPs abundance and wind direction [36]. However, research conducted in urbanized regions established that the concentration of atmospheric MPs was considerably associated with the direction of the wind [8,49]. Ryan et al. (2023) collected atmospheric fallout samples during Hurricane Larry as it traversed the remote area of Newfoundland (Canada) in September 2021. MPs were detected in all samples taken before, during, and after the storm. At the peak of the storm, there was the highest level of 113,569 ± 29.215 MPs m⁻² day⁻¹, followed by a decline in deposition after the storm had passed. Furthermore, MPs deposition exhibited a relatively similar pattern in both pre- and post-hurricane samples. However, the total particle count was slightly higher, on average, after the hurricane compared to before [54]. These findings demonstrate that intense atmospheric events can deposit MPs in remote locations where typically low levels of MPs deposition would be expected. Furthermore, Allen et al. (2019) discovered a negative correlation between MP concentrations in the atmosphere and both rainfall and snowfall. They suggest that the increase in frequency and severity of these weather events, as opposed to their length, could enhance MPs deposition [34].

3.1.3. Fates of Deposited MPs

After being deposited into terrestrial or aquatic environments, MPs can become airborne again through resuspension. Precisely, the movement of air caused by wind, in fact, can result in a release of MP particles back into the atmosphere. Moreover, windy weather allows some MPs initially released on the ground to enter the atmospheric environment, bringing in additional MPs from external sources and facilitating their airborne dispersion [53]. MPs might be also released from the seawater into the atmosphere by sea-spray and bubbles bursting [36,61]. Bubble bursting is a phenomenon that takes place at the air–sea interface and determines the generation of tiny particles that originate from the surface of the ocean and are discharged into the atmosphere [42]. However, this mechanism needs further research. The dispersion of MP particles is also influenced by their physical–chemical properties. The aerodynamic diameter of particles, which is a function of their density, size, and aspect ratio, plays a crucial role in determining their atmospheric residence time, deposition velocity, and re-suspension [53]. The findings suggest that particle size may not be a decisive factor in the deposition and suspension of atmospheric MPs. It suggests that fragment-shaped particles are more prone to deposition compared to fibers. Denser particles tend to settle more rapidly, while less dense ones are more susceptible to being lifted from the ground by the wind [39,53].

Therefore, MPs can be emitted from different sources and end up in the atmosphere [8,38], where they can be blown away by the wind [1,37,59] and then removed from the atmosphere by wet or dry deposition and end up in the aquatic and terrestrial environment. However, once the MPs are deposited, they can be resuspended in the atmosphere, thus closing the so-called atmospheric plastic cycle [5,9].

3.2. Analytical Methodologies

To assess the implications of MPs presence in the atmosphere in depth, reliable and advanced analytical techniques capable of accurately measuring and identifying the MPs in the air are a matter of urgency.

3.2.1. Sampling Methods

Regarding the sampling activities, active [2,7,9,13,20,21,23,36,42,43,46,47,48,52,55,59,61,64,68] and passive sampling [8,18,22,23,25,26,31,32,33,39,40,45,48,49,51,53,54,57,65] of MPs in the atmosphere are the two approaches used to collect samples of plastic particles in the air. Passive sampling provides data on wet deposition (i.e., collection of precipitation), dry deposition (i.e., in the absence of precipitation), or bulk deposition (wet and dry deposition) of MPs present in the atmosphere. This sampling is conducted using aseptic metallic or glass containers (e.g., passive funnel or beaker-type samplers) that are usually placed on the rooftop of a building in the sampling area [32,33]. The benefits of these samplers are their ease of use and no requirement for power to the study site, allowing studies in remote locations with minimal infrastructure at a very low cost. Nevertheless, the disadvantages of these methods are sample resuspension via wind or evaporation, the non-standardized collection equipment, and the collection of a range of wet and/or dry deposition for varying periods and precipitation quantities [18,30,40]. Knobloch et al. (2021) tested four passive deposition samplers: a glass bottle with a funnel, an open glass beaker, a Petri dish covered in double-sided polyethylene adhesive tape, and an N-CON ADS/NTN Atmospheric Precipitation Sampler [18]. The adhesive sampler was discarded, as no particles could be removed from the tape.

The N-CON ADS/NTN Atmospheric Precipitation Sampler had a heart attack system that determined the opening and closing of the container in the presence and absence of rain, allowing the collection of wet deposition. This method prevented resuspension of MPs, but the authors were not able to draw solid conclusions about the sampling effectiveness. Both the bottle with the funnel and beaker showed effective dry and wet deposition sampling for microplastics, concluding that either design is suitable [18]. Active sampling is an effective method of sampling atmospheric microplastics. It consists in the use of pumped samplers of known volumes of air in defined periods and selected locations. In this way, MPs present in the air are collected on filters and then analyzed [2,9,46,68]. The most often used filters are quartz fiber filters [2,6,9,45,46,47,48,52,68] or fiberglass filters [4,7,13,15,35,36,43,46,50,53,54,55,56,59]. Other types of filters used are silver membrane filters [1,5,15], cellulose acetate filters [30,49], alumina-based membrane filters [5], and nitrocellulose membrane filters [51].

Thus, active samplers collect air samples and allow us to measure MPs that are currently present in the air mass, rather than those that have settled as pollution (as in the case of passive samplers). By actively pumping air, these samplers estimate the amount of airborne MPs that have not yet been deposited. Therefore, it is suggested to use passive samplers to gather atmospheric deposition (both wet and dry), alongside active pumped-air sampling, to obtain a comprehensive understanding of the microplastic content in the air.

3.2.2. Analytical Measurements

Visual observations, µ-Fourier transform infrared spectroscopy (µ-FTIR), and µ-Raman spectroscopy are commonly used techniques to analyze (Table 3) (i.e., quali-quantification) MPs in the atmosphere [2]. Visual observations are generally used for the screening and quantification of MPs and can be conducted using various devices [5]. Among the studies considered in this review, 21 used a stereomicroscope for the characterization of physical characteristics and quantification of MPs, with optical microscope (n = 10), scanning electron microscope (SEM) (n = 10), and fluorescence microscope (n = 8) being much less common. Visual methods are generally effective for identifying and quantifying MP particles, as well as the sizes, shapes, and colors. However, such methods cannot identify the type of polymer or conclusively confirm that a particle is plastic. Moreover, visual identification is subject to human bias and dimensional limitations due to the microscope’s resolution [16]. More recent studies combine visual identification with a chemical analysis method to confirm the presence of MPs [17,42,47].

Vibrational spectroscopic methods, such as µ-Fourier transform infrared (FTIR) or µ-Raman spectroscopy, are widely used for the identification and quantification of MPs. These techniques offer non-destructive approaches to identify plastic polymeric particles based on the characteristic spectra of the molecular vibration fingerprints. By examining the characteristic spectra of MP samples, researchers can determine their chemical composition and structural information and the comparison of unknown MP spectra with known reference spectra aids in accurately recognizing and categorizing plastic particles [69]. Dris et al. (2016) showed the first application of µ-Fourier transform infrared (µ-FTIR) micro spectroscopy coupled with an attenuated total reflectance (ATR) accessory for the chemical characterization of fiber in the atmospheric fallout to estimate the proportion of synthetic and natural fibers [46]. However, some disadvantages associated with µ-FTIR include the cost of the instrument and the capability of the operator as well as the thickness and the shape of the sample and the detection limit for the size that varies among different instruments. µ-ATR–FTIR is suitable for larger particles (typically larger than 500 μm), while µ-FTIR coupled with microscopy (μ-FTIR) can be used for particles up to 20 μm [69,70]. Unlike µ-FTIR spectroscopy, µ-Raman spectroscopy is not constrained by the thickness or shape of the sample and is suitable for wet samples, while µ-FTIR is influenced by the presence of water because it is active in the infrared region. µ-Raman spectroscopy allows obtaining spectra based on the interaction between laser light and sample molecules and can detect smaller particles up to 1 μm for μ-Raman techniques. However, µ-Raman spectroscopy is subject to interference from compounds that produce fluorescence, and therefore requires a more rigorous removal of organic matter to minimize errors [71].

To investigate atmospheric MP pollution in the catchment of the Weser River, Kerchen et al. (2022) identified all MP particles sampled using µ-FTIR with a focal plane array (FPA) for MP particles > 11 µm and µ-Raman spectroscopy equipped with a 532 nm laser and a CCD for MP particles < 4 µm [30]. Pyrolysis coupled with mass spectrometry (Py-GC–MS) is a non-visual method for the identification of MPs in samples [38]. Pyrolysis allows the identification of the plastic polymer and its concentration in the sample. This methodology usually does not require sample preparation for airborne MPs and is not influenced by the size and shape of the test samples [7]. However, it is not possible to define the number and shape of particles, as it is a destructive technique [24], and it may be not representative due to the limited amount of sample that can be used for the analysis [7]. Thus far, only a few studies have used pyrolysis for the identification of MPs in the atmosphere. O’Brien et al. (2021) subjected samples of road dust to double-shot Py-GC–MS analysis that allows identification of PVC (29%) and PET (29%), which where the most abundant MPs detected within the total quantified mass of MPs from all samples (mg g⁻¹), followed by PE (21%), PP (10%), poly (methyl methacrylate) (7.6%), and PS (1.1%) [22].

Table 3. Summary of analytical measuring techniques used by the included studies for the identification and quantification of atmospheric MPs.

Technique	Size Limitation of the Technique	Microplastics Size Range	Ref.
μ-Raman	>0.45 μm	<100 to >1000 μm	[39]
μ-Raman	>0.45 μm	<2.5 μm	[6]
μ-Raman	>0.45 μm	<2.5 to 2.600 μm	[34]
μ-Raman	>0.45 μm	1.60 to 16.80 μm	[4]
μ-ATR–FTIR	>7 μm	50.01 to 1579.43 μm	[40]
μ-ATR–FTIR	>7 μm	>50 μm	[60]
μ-Raman	>0.45 μm	<100 μm	[41]
μ-FTIR	>20 μm	140 to 9.96 mm	[19]
μ-FTIR	>20 μm	<200 to >4200 μm	[33]
μ-FTIR SEM	>20 μm N/A	<10 μm	[42]
μ-ATR–FTIR	>7 μm	N/A	[43]
μ-Raman	>0.45 μm	5 to 5000 μm	[2]
μ-ATR–FTIR	>7 μm	<50 to 200 μm	[1]
μ-Raman	>0.45 μm	10 to 3094 μm	[44]
μ-ATR–FTIR	>7 μm	10 to 4556 μm	[36]
μ-FTIR	>20 μm	50 to 2210 μm	[35]
Stereomicroscope	>100 μm	100 to 5000 μm	[32]
μ-ATR–FTIR	>7 μm	50 to 5000 μm	[45]
μ-ATR–FTIR	>7 μm	50 to 4850 μm	[46]
μ-Raman	>0.45 μm	158 to 509 μm	[37]
Py-GC/MS μ-FTIR	0.01–4 μg >20 μm	N/A	[47]
μ-ATR–FTIR Py-GC/MS	>7 μm	NA	[26]
μ-FTIR	>20 μm	40 to 800 μm	[48]
μ-ATR–FTIR μ-Raman	>7 μm >0.45 μm	11 to 1945 μm	[30]
μ-Raman	>0.45 μm	<63 to 5000 μm	[49]
μ-FTIR	>20 μm	<200 to >500 μm	[18]
μ-FTIR	>20 μm	>20 μm	[72]
μ-FTIR	>20 μm	16.14 to 2500 μm	[13]
μ-ATR–FTIR	>7 μm	20 to 2000 μm	[25]
LDIR	>20 μm	37.7 to 95.79 μm	[8]
μ-FTIR	>20 μm	50 to 500 μm	[50]
μ-Raman	>0.45 μm	>2.5 μm	[15]
μ-FTIR	>20 μm	<35 to >50 μm	[23]
μ-FTIR	>20 μm	<500 to >4000 μm	[51]
μ-Raman	>0.45 μm	<5 to >30 μm	[68]
μ-Raman	>0.45 μm	183 to 11,877 μm	[52]
Py-GC/MS	N/A	<250 to 5000 μm	[22]
μ-FTIR	>20 μm	>20 μm	[43]
μ-Raman	>0.45 μm	2 to 100 μm	[54]
μ-FTIR	>20 μm	>20 to 23,565μm	[20]
μ-Raman	>0.45 μm	0.45 to 2800 μm	[64]
μ-FTIR	>20 μm	35.38 to 1378.89 μm	[55]
μ-FTIR	>20 μm	58.59 to 2251.54 μm	[59]
μ-FTIR	>20 μm	25 to 3000 μm	[5]
μ-Raman	>0.45 μm	2.40 to 2181.48 μm	[73]
μ-FTIR	>20 μm	12.96 to 333.62 μm	[56]
μ-Raman	>0.45 μm	<500 μm	[57]

Table 3. Summary of analytical measuring techniques used by the included studies for the identification and quantification of atmospheric MPs.

Technique	Size Limitation of the Technique	Microplastics Size Range	Ref.
μ-Raman	>0.45 μm	<100 to >1000 μm	[39]
μ-Raman	>0.45 μm	<2.5 μm	[6]
μ-Raman	>0.45 μm	<2.5 to 2.600 μm	[34]
μ-Raman	>0.45 μm	1.60 to 16.80 μm	[4]
μ-ATR–FTIR	>7 μm	50.01 to 1579.43 μm	[40]
μ-ATR–FTIR	>7 μm	>50 μm	[60]
μ-Raman	>0.45 μm	<100 μm	[41]
μ-FTIR	>20 μm	140 to 9.96 mm	[19]
μ-FTIR	>20 μm	<200 to >4200 μm	[33]
μ-FTIR SEM	>20 μm N/A	<10 μm	[42]
μ-ATR–FTIR	>7 μm	N/A	[43]
μ-Raman	>0.45 μm	5 to 5000 μm	[2]
μ-ATR–FTIR	>7 μm	<50 to 200 μm	[1]
μ-Raman	>0.45 μm	10 to 3094 μm	[44]
μ-ATR–FTIR	>7 μm	10 to 4556 μm	[36]
μ-FTIR	>20 μm	50 to 2210 μm	[35]
Stereomicroscope	>100 μm	100 to 5000 μm	[32]
μ-ATR–FTIR	>7 μm	50 to 5000 μm	[45]
μ-ATR–FTIR	>7 μm	50 to 4850 μm	[46]
μ-Raman	>0.45 μm	158 to 509 μm	[37]
Py-GC/MS μ-FTIR	0.01–4 μg >20 μm	N/A	[47]
μ-ATR–FTIR Py-GC/MS	>7 μm	NA	[26]
μ-FTIR	>20 μm	40 to 800 μm	[48]
μ-ATR–FTIR μ-Raman	>7 μm >0.45 μm	11 to 1945 μm	[30]
μ-Raman	>0.45 μm	<63 to 5000 μm	[49]
μ-FTIR	>20 μm	<200 to >500 μm	[18]
μ-FTIR	>20 μm	>20 μm	[72]
μ-FTIR	>20 μm	16.14 to 2500 μm	[13]
μ-ATR–FTIR	>7 μm	20 to 2000 μm	[25]
LDIR	>20 μm	37.7 to 95.79 μm	[8]
μ-FTIR	>20 μm	50 to 500 μm	[50]
μ-Raman	>0.45 μm	>2.5 μm	[15]
μ-FTIR	>20 μm	<35 to >50 μm	[23]
μ-FTIR	>20 μm	<500 to >4000 μm	[51]
μ-Raman	>0.45 μm	<5 to >30 μm	[68]
μ-Raman	>0.45 μm	183 to 11,877 μm	[52]
Py-GC/MS	N/A	<250 to 5000 μm	[22]
μ-FTIR	>20 μm	>20 μm	[43]
μ-Raman	>0.45 μm	2 to 100 μm	[54]
μ-FTIR	>20 μm	>20 to 23,565μm	[20]
μ-Raman	>0.45 μm	0.45 to 2800 μm	[64]
μ-FTIR	>20 μm	35.38 to 1378.89 μm	[55]
μ-FTIR	>20 μm	58.59 to 2251.54 μm	[59]
μ-FTIR	>20 μm	25 to 3000 μm	[5]
μ-Raman	>0.45 μm	2.40 to 2181.48 μm	[73]
μ-FTIR	>20 μm	12.96 to 333.62 μm	[56]
μ-Raman	>0.45 μm	<500 μm	[57]

Overall Summary

The most commonly used visual methods for the analysis of airborne MPs include µ-FTIR spectroscopy and µ-Raman spectroscopy. µ-FTIR spectroscopy is effective for particles larger than 20 µm, whereas µ-Raman spectroscopy is suitable for particles larger than 0.45 µm. These techniques enable the identification and quantification of MPs but do not allow for the determination of polymer type. They utilize non-destructive approaches to identify plastic polymeric particles by analyzing their characteristic molecular vibration spectra. However, certain aspects such as the shape, thickness, and color of microplastics, the presence of water, and compounds producing fluorescence in the sample can cause interference with the analysis. A non-visual technique, unaffected by the aspects mentioned earlier, is Py-GC–MS, which allows for the identification of plastic polymers and their concentration in the sample. However, it is a destructive technique that may be non-representative due to the limited sample quantity available for analysis and does not allow for defining the number and shape of particles.

3.3. Health Implications

MP particles are ubiquitously present in the air, particularly microfibers smaller than 1650 µm [46]. Researchers have detected MPs in both indoor and outdoor air samples from various locations worldwide, revealing a higher abundance of MPs in indoor compared to outdoor environments [5,38].

When humans inhale aerosol particles, some of them are retained within the respiratory tract, potentially depositing on mucous membranes and alveolar surfaces, whereas others are expelled through exhalation [30]. As humans take between 10 and 20 breaths per minute at rest, airborne MPs can easily be inhaled and enter the body through the respiratory system [6]. The toxic effects of MPs in the lungs primarily manifest as the promotion of oxidative stress, apoptosis, and the induction of pulmonary fibrosis [74]. The impact of MPs is dependent on various factors such as particle size, burden, density, charge, type, and individual factors. Furthermore, the environmental changes that occur during MPs weathering also play a significant role in determining the biological impact of MPs on the lungs. Currently, there is a lack of significant in vitro or in vivo data regarding the potential toxicity of MP particles or fibers in human lung cells. This information is crucial, as weathering of MPs in the environment seems to promote the cellular internalization of MPs [4]. As the environment continuously processes MPs, they undergo weathering and fragmentation, resulting in increasingly smaller pieces. This increases the likelihood of inhalation into deeper parts of the lungs and the potential for negative health impacts [40].

The presence of MPs in human lung tissues obtained at autopsy was characterized for the first time by Amato-Lourenço et al. (2021). In 13 of the 20 tissue samples, 31 synthetic polymer particles (sizes ranging from 1.60 to 5.56 μm) and fibers (sizes ranging from 8.12 to 16.8 μm) were observed, and PP and PE were the most common particles found. It is worthy to underline that the limited number sample cases studied is small to identify trustworthy associations or to elucidate the potential adverse health effects of these compounds [4]. Instead, Baeza-Martinez et al. (2022) investigated the presence of MPs in the lower human airways using bronchoalveolar lavage fluid (i.e., BALF). Results of their study indicate that BALF samples were contaminated with an average microfiber concentration of 9.18 ± 2.45 items 100 mL⁻¹ BALF and with an average size of 1.73 ± 0.15 mm [19]. Danso et al. (2022) investigated the lung toxicity response to PS, PP, and PVC MPs in three strains of mice (i.e., C57BL/6, BALB/c, and ICR) by instilling them with 5 mg kg⁻¹ of the three MPs in 50 µL of saline solution for two weeks. The mice strains had different genetic backgrounds that caused different immune response processes. According to this, authors hypothesized that the genetic background could potentially influence the lung immune response to MP exposures. Their results indicated that there was an important increase in the number of total cells, macrophages, eosinophils, neutrophils, and lymphocytes in the bronchoalveolar lavage fluid (BALF) of mice (C57BL/6) instilled with PS microplastics compared to those in the control groups. Additionally, in the BALS of PP-instilled ICR mice, there was also an increase in inflammatory cells (i.e., total cells, macrophages, eosinophils, neutrophils) compared to the control group. However, exposure to PVC did not induce any important changes in the inflammatory cellular level across the three strains of mice. Notably, BALB/c mice exhibited no significant inflammatory responses following stimulations with the three MPs (i.e., PS, PP, PVC) [14]. Instead, Zha et al. (2023) examined the influence of airborne MPs and nanoplastics (NPs) on the nasal and lung microbiota in mice through a series of bioinformatic and statistical analyses. Mice were allocated into three groups, the MP group, the NP group, and a negative control group. Each individual was exposed to 5 μm PS MPs, 99 nm PS NPs, both at concentrations of 10 μg μL⁻¹, and sterile saline, respectively. According to PERMANOVA, NMDS, and PCoA results, both MPs and NPs have the potential to impact various bacteria in the respiratory tract, leading to microbial ecological imbalances in the nasal cavity and lungs. Moreover, results revealed that the lung composition was similar between the NP and NC groups, but different between the MP and NC groups. This suggests that MPs have a stronger influence on lung microbiota than NPs [75].

It was suggested that MPs inhalation and ingestion can lead to tissue toxicity, as well as inflammation and infections resulting from the absorption of toxic chemicals and microorganisms present on the MPs. Amato-Lourenco et al. (2022) provides the first evidence of a correlation between the amount of SARS-CoV-2 in the air and the presence of MP fibers. This paper suggests that SARS-CoV-2 has the potential to attach to MPs via adsorption. This could potentially facilitate the transmission of the virus to the human body due to the extensive surface area of MP fibers, which could act as carriers for the virus [40]. Furthermore, some studies indicate that MPs have a propensity to adsorb and amass pollutants, including heavy metals, polychlorinated biphenyls (PCBs), and polycyclic aromatic hydrocarbons (PAHs). As a result of their widespread distribution and potential as mutagens and carcinogens, there has been an escalating public concern over this issue [1,76].

4. Discussion

Current investigations about atmospheric MPs primarily concentrate on urban, suburban, and remote regions, with limited importance on farmland ecosystems. In farmlands, atmospheric deposition stands out as a significant contributor to the presence of MPs. Crops could play a key role in the uptake of MPs through roots and leaves (<100 µm) [77]. This process sets in motion a potential transfer through the food chain, amplifying the risk of human exposure to MPs [51]. This highlights the significance of enhancing the study of factors influencing the atmospheric deposition of MPs. Beyond traditional considerations, such as source emission and spatial distribution, it becomes essential to incorporate into research additional factors in the analysis of airborne MPs, such as natural or anthropic phenomena that can cause deposition or resuspension of atmospheric MPs. For instance, during transport, certain microplastics may be deposited on the leaves of plants and trees, establishing them as temporary sinks for MPs. To explore this phenomenon, some studies have employed leaves as passive samplers for the biomonitoring of MPs [72,78]. In their study, Leonard et al. (2023) aimed to identify factors influencing the concentration of MPs on leaves. The considered factors included leaf position above the ground, land-use type, and leaf surface hydrophobicity. Notably, among such factors, leaf surface hydrophobicity emerged as the most significant contributor to the variability in MPs concentration [72]. The aim of Liu et al.’s (2020) research, instead, was to gain insights into the environmental behavior of atmospheric MPs during their movement and to create a dataset for flux estimation. However, the presence of MPs on leaves may exhibit greater dynamism, with storage potentially being of a temporary nature [78]. The results of this study imply that leaves may not function as a consistently reliable passive sampler, unless the sampling protocol is adjusted to accommodate uncertainties related to concentration variations [72,78].

Moreover, accumulation of MPs in human lung tissues needs to be clarified. MPs exposure to humans can occur predominantly through ingestion and inhalation [79]. The ingestion of MPs is supported by various studies; however, knowledge regarding inhalation is still lacking. This is attributed to the fact that only recently has evidence of the presence of MPs in atmospheric fallout been increasing [33]. The majority of MPs found in the atmosphere are made of fibers of various lengths, presumably because of their favorable aerodynamic properties. Their abundances are predominantly higher in urban areas with high population density [39,45]. This highlights the potential for human exposure to MPs through inhalation. However, due to analytical difficulties, studies on the accumulation of MPs in the human body are still limited. Furthermore, it is important to question the existing studies’ findings regarding the likelihood of MPs translocation based on particle size [70].

It has been estimated that an average male engaging in light activity could potentially inhale up to 272 MPs particles day⁻¹. However, recent research has found only approximately 0.56 MP particles g⁻¹ of lung tissue, suggesting that a large percentage of inhaled MPs are retained in the upper airways [4]. Depending on their size, such particles can deposit on respiratory tract mucous membranes and alveolar surfaces. The smaller the particles are, the deeper they can penetrate the respiratory tract and be retained for longer periods: particles smaller than 100 μm can enter the upper respiratory system through the nose and mouth; particles smaller than 10 μm have the potential to reach the lungs and upper bronchial airways; and those with a diameter smaller than 2.5 μm can settle in the lower bronchial airways and alveolar region and can even reach the gas exchange area in the lungs. From a toxicological standpoint, any particle with a diameter smaller than 10 μm has the potential to cause biological activity in susceptible individuals, and only particles less than 2.5 μm have the potential to enter the circulatory system [4,30]. The shape and size of microplastics influence the potential health risk they pose.

When microplastics less than 10 µm are inhaled, they can accumulate in the lung and cause localized particle toxicity, triggering an immune response. Once inside the respiratory tract, most MPs are trapped in the fluid lining the lungs. However, certain fibers may avoid the lung’s natural clearing mechanisms and lead to inflammation, either temporarily or over a long period. Particle size also plays a role in toxicity, with thinner fibers being breathable and longer ones posing more persistence and toxicity to lung cells [80]. As reported by Wright at al. (2017), fibers with diameters of 15–20 µm are not efficiently cleared by alveolar macrophages and the mucociliary escalator in the lungs. On the other hand, fibers that are less than 0.3 μm thick but greater than 10 μm long are more carcinogenic because they can bypass the body’s natural defense mechanisms and remain in the lungs for longer periods of time, increasing the risk of lung cancer [80]. For this reason, it is possible to suggest that the identification of fibers with an average size of 1.73 ± 0.15 mm in the study conducted by Baeza-Martinez et al. (2022) may be attributed to methodological artifacts.

Such artifacts are due to the absence of standardized analysis procedures and standardized reference materials that ensure the absence of plastics and other contaminants in the chemicals used for sample analysis (even if declared as high quality). Regarding the analysis of airborne microplastics, after samples collection, usually pretreatment is not required. The samples are directly subjected to chemical analysis (e.g., spectrophotometric techniques, pyrolytic techniques). However, for more complex matrices, such as food matrices, sample pretreatment with alkaline digestion methods is necessary for the removal of the organic matrix. This pretreatment is then followed by filtration (with filter membranes of various sizes (0.22–20 µm) and materials (e.g., cellulose nitrate, glass fiber)) and subsequent chemical analysis. Regarding chemical analysis, spectrophotometric techniques allow for the qualitative and quantitative analysis of MPs but do not enable the determination of the polymer type. On the other hand, analysis using Py-GC–MS allows for the identification of the polymer type and its concentration in the sample but does not provide information on the number, shape, size, and color of the particles. Therefore, a synergistic approach combining both techniques allows for a comprehensive analysis of MPs. Each step in the preparation and every material in contact with the sample can result in external contamination. For this reason, the use of procedural blanks is fundamentally important. However, the exclusive reliance on procedural blanks does not guarantee the preservation of measurement accuracy. Methodological artifacts may arise not only from procedural blanks but also from human errors and/or inadequate quality assurance protocols. It is important to implement rigorous quality assurance and quality control (QA/QC) measures. These protocols are crucial for preventing any unintended extraneous or background contamination in the samples, ensuring the high quality and quantity of data obtained. Currently, no guidelines or protocols have been published for QA/QC [81].

Furthermore, the scarcity of research investigating the accumulation of MPs in the human body underscores the challenges associated with accurate analysis. Hence, to validate the existence of MPs in the human body, it is imperative to conduct further investigations using more rigorous methodologies.

5. Conclusions

Thanks to the analysis of various studies in the literature, it is possible to say that MPs are ubiquitous emerging contaminants. MPs of different shapes (e.g., fiber and fragment) and sizes (0.1 µm to 5000 µm) have been identified, both indoors and outdoors. However, drawing conclusions about the effects of inhaled microplastics on human health is difficult due to the heterogeneity of the results reported by different studies and the lack of standardization of analytical methods. In the literature, there are several instrumental (i.e., interference) and procedural (i.e., methodological artifacts) limitations. Moreover, the sources of these pollutants and the factors involved in their distribution are still unclear. Finally, the effects of inhaling microplastics on human health must be well documented in order to develop prevention and treatment measures in the future. For these reasons, validated methods is a challenge yet to be achieved that requires a synergistic effort by the scientific community and authorities to study and identify standardized methods for sampling and measuring microplastics.