Abstract
This paper presents an overview of the literature studies on the sources of ultrafine particles (UFPs), nanomaterials (NMs), and nanoparticles (NPs) occurring in indoor (occupational and residential) and outdoor environments. Information on the relevant emission factors, particle concentrations, size, and compositions is provided, and health relevance of UFPs and NPs is discussed. Particular attention is focused on the fraction of particles that upon inhalation deposit on the olfactory bulb, because these particles can possibly translocate to brain and their possible role in neurodegenerative diseases is an important issue emerging in the recent literature.
INTRODUCTION
Ultrafine particles (UFPs) are defined as particles with diameter below 100 nm. Usually reference is made to the electrical mobility diameter, since the instruments adopted to measure their concentration rely upon their electrical mobility. An estimate of the amount of their emissions, limited to the European Union for 2008, is of 271 ktonnes, of which 5%, 34%, 22%, 15%, 12%, 4%, and 8%, respectively, are due to industrial processes, road transport, other transport and machinery, residential and commercial, industrial combustion, power generation, and agriculture sources [1].
UFPs have drawn the attention of the scientific community in the last few decades because their causative role has been demonstrated in important respiratory and cardiovascular pathologies. Traditionally, the health effects of airborne aerosols have been put in relation with mass-based reference limit values both in outdoor, such as particulate matter (PM) of 10 μm or less in diameter (PM10) and PM2.5 reference values and in indoor environments, specifically in occupational settings where Threshold Limit Value (TLV) or Occupational Exposure Limit value (OEL) are applied. However, mass-based aerosol concentrations underrepresent the contribution of UFPs, due to their negligible mass in comparison with coarse particles. On the contrary, UFPs are abundantly present in both indoor and outdoor environments, so that it is their number concentration that gives a proper indication of their presence. These particles may elicit their effects within cellular targets, due to their dimensions comparable with those of the cellular structures. Moreover, their small size allows them to penetrate into the respiratory system with high efficiency down to the alveolar region and to be distributed through the blood circulation to peripheral organs. Another important health-relevant physical property of these particles is their surface area. The importance of this parameter is due to the ability to adsorb noxious substances and to vehiculate them into the body. Moreover, their ability to generate reactive oxygen species (ROS) have been explained through their higher surface area per unit mass and to their surface reactivity [2].
An important feature of UFPs is their fast evolution, particularly of their smallest fractions below 10–20 nm. These particles have very high Brownian diffusion, so that they easily impact on other particles and on available surfaces; consequently they have very short atmospheric lifetimes [3]. As a result of that, their concentrations quickly decay with increasing distances from the emission sources and yet being in proximity of them, with time, in the scale of few seconds, because the coagulation mechanism is very efficient due to the high particle concentration in proximity of the sources and then to the high frequency of impaction on other particles. In addition, particle growth due to semivolatile compounds condensation on their surfaces also occurs [4]. This feature has two very important implications. The first is that individual exposures to UFPs should be assessed in the microenvironments where individuals actually reside or pass through, rather than relying on measurements performed at fixed monitoring stations. The second is that the time resolution of their measurement should be in the time scale of their evolution to capture the transient spike emissions from their sources [5].
Moreover, particles of the same size of UFPs not only may be released unintentionally, because of the chemical and physical transformations that materials undergo, they may be introduced also intentionally to take advantage of their specific nanoscale properties. In such case, they are referred to as nanomaterials (NMs) and nanoparticles (NPs). The European Commission SCENIHR [6] defines nanomaterials and nanoparticles, respectively, as “Any form of a material that is composed of discrete functional parts, many of which have one or more dimensions of the order of 100 nm or less.” and “A discrete entity which has three dimensions of the order of 100 nm or less.”
Within this context, the present narrative review summarizes the main UFP, NM, and NP emission sources occurring in indoor (occupational and residential) and outdoor environments.
OUTDOOR SOURCES
Natural sources
The order of magnitude of particle number (PN) concentrations in uncontaminated outdoor environments is of about some hundreds particles cm–3. As found in the Amazon tropical forest, median values of about 400 particles cm–3 and 600 particles cm–3 have been reported during the rain and the dry seasons, respectively [7].
PN concentrations may be significantly increased above such low levels, due the formation of new particles in atmosphere. This is a widely studied process because atmospheric particles have strong influences on the microphysics of clouds and, then, on the rainfalls and the global climate [8].
The formation of new particles proceeds through a nucleation process involving the condensation of supersaturated molecular species into clusters (below 1 nm in size) that grow into stable larger particles. In this process, there is evidence that sulphuric acid, ammonia, and low volatility oxidized organic compounds are involved [9]. As to the latter, forests may play an important role due to their abundant emission of non-methane hydrocarbon, such as terpenes, that, once oxidized, may condense to form organic aerosols [10, 11]. For instance, in the Southern Finland boreal forest, during intense nucleation and growth events, PN concentration increasing from about 3×103 particles cm–3 to above 2×104 particles cm–3 have been reported [12].
Biogenic marine emissions as well may be relevant in sites affected by marine air masses. New particles may nucleate from condensable iodine vapors deriving from the photolysis of biogenic hydrocarbons emitted by algae [13]. It has also been proposed that new particles are formed starting from dimethylsulphide, released by planktonic algae in sea water, after oxidation in the atmosphere to form sulphate aerosol [14].
Another important natural sources of UFPs are represented by forest fires and volcano eruptions [15].
Guyon et al. [16] based on 69 plumes from biomass burning in Amazonia, sampled within the boundary layer, estimated an aerosol geometric diameter of 110±15 nm, with emission factors ranging from 2.3×1014 to 5.4×1015 particles emitted per kg dry matter burnt.
Following a series of eruptions of the Eyjafjallajökull volcano in Iceland started on April 2010, Schäfer et al. [17] observed peak UFP concentrations during a period of intense vertical atmospheric mixing and advection followed by stable atmospheric conditions in southern Germany. Particle concentrations in the range 10–100 nm as high as 16000 particles cm–3, compared with considerably lower background values (<4000 particles cm–3) were observed with a temporal trend highly correlated with sulphur dioxide (SO2) transported in the volcanic ash plumes [18]. The authors concluded that the increase of UFP concentration was due to the high concentrations of sulphuric acid formed by the photochemical oxidation of SO2 in the daytime.
Industrial emissions
The number concentration and the size distributions of UFPs at the stacks of industrial plants are strictly dependent on the materials processed and on the kind of transformations that they undergo, as well as on the systems adopted to abate pollutant emissions. Aerosol emitted by industrial sources is made up by two main modes, according to their aerodynamic diameter (da): a coarse mode (da > 2.5 μm) and a fine one (da < 2.5 μm), that include UFPs and accumulation mode particles (0.1–1 μm). Due to their high deposition rates, coarse particles efficiently deposit to the ground within short distances from the emission point, whereas fine particles may travel for longer distances [19]. This means that they may elicit their health effects within long reach from the emission point. Once emitted, aerosol particles quickly undergo rapid transformations that affect their concentration and size distributions. The plume dispersion, and hence its concentration, depends on a number of factors such as stack gas temperature, stack gas velocity, stack height, heat emission rate, atmospheric wind speed, and atmospheric stability. The particle size distribution will evolve over time due to the coagulation process, favored by the high particle concentrations in the emission plumes, as well as due to the condensation of supersaturated vapors on the surface of existing particles. The latter is a process occurring in completion with the nucleation of new particles from the gas phase.
Steel factory emissions
Leoni et al. [20] measured UFP peak concentration as high as 3.2×105 particles cm–3 at 1.5 km downwind from an iron and steel factory. In the plume, UFPs were mostly present in the range 19–44 nm in diameter. Their average content in polycyclic aromatic hydrocarbons (PAHs) was 2.1 mg g–1 and 0.9 mg g–1 of seven carcinogenic PAHs were on average present in the quasi-ultrafine size fraction (da < 170 nm).
PM2.5 concentrations measured at the stack emissions of a sinter steelmaking plant in UK varied between 8.5 and 25.24 mg m–3 and were made up by 90% PM1 and from 11% to 39% of PM0.1. The major mode diameters were at 120 nm, both for the mass and the number metrics. Particles with sizes below 120 nm represented 39% and 73%, respectively, of the mass and of the number of particles of the aerosol sampled, whereas 11% and 28%, respectively, in terms of mass and of number were due to particle with diameters below 70 nm. Metals and anions amounted respectively to about 39% and 4% of the ultrafine fraction. Elements such as Na, Pb, Cu, and Al present at concentrations from 20 to 500 μg m–3; Cd, Se, Zn, and Th present at concentrations from 2 to 100 μg m–3; As and Ag present at concentration from 0.2 to 2 μg m–3, displayed size mass distribution with main mode slightly above 120 nm [21].
The total concentration of the 16 U.S. EPA priority pollutant PAHs [22] and of the International Agency for Research on Cancer (IARC) group 1 carcinogenic Benzo[a]Pyrene (B[a]P) [23] were, respectively, of about 17 μg m–3 and 1.9 μg m–3. 61% and 34% of PAHs were, respectively, found in the fine and ultrafine fractions.
PM2.5 concentrations in fugitive emissions from a basic oxygen steelmaking plant were extremely variable, ranging from 0.5 mg m–3 to 20–35 mg m–3, 50% –60% of them was made up of PM1 and 2–3% of PM0.1. Particle concentrations of about 4×105–1.1×106 particles cm–3 were reported, with a major mode in the ultrafine region. The mass contribution of metals such as Fe, Mn (>500 μg m–3), Na, Mg, K, Zn, Pb, Ca (11–20 μg m–3), Cr, Ni, Cu, Al (2–10 μg m–3), Cd, Co, As, V, Ba, Se, Sr (0.03–2 μg m–3) represented 29–40% and 47–57%, respectively, of the ultrafine and fine fractions. The US EPA 16 PAH and the B[a]P concentrations ranged from 88 to 154 ng m–3 and from 3.8 to 7.9 ng m–3, respectively. PAHs were mainly present in the fine (64%) and ultrafine (29%) fractions [21].
Aerosol measurements carried out 30–60 m far from a French coke oven plant revealed PN concentrations extremely variable, from about 5,000 particles cm–3 to 45,000 particles cm–3, with peak concentrations presumably due to coke oven fugitive emissions. An ultrafine mode, due to oven emissions, was present with concentrations above 10,000 particles cm–3. Metals such as Fe (concentrations from about 4.5 to 14.0 μg m–3), Al (0.7–1.2 μg m–3), Cr (0.5–1.4 μg m–3), Mn (0.3–1.8 μg m–3), Ni (1.1–2.0 μg m–3), and Cu (of about 0.3 μg m–3) were mainly distributed in the fine fraction. The total US EPA 16 PAH and B[a]P concentrations ranged respectively from 198 to 2,465 ng m–3 and from 8.8 to 301 ng m–3. Oven door leakages were responsible for the highest concentrations measured. 77%, 14%, and 9% of PAHs were distributed in the fine, ultrafine, and coarse fractions, respectively [21].
Waste incineration plants
Buonanno et al. [24] measured total PN concentrations between 1×105 particles cm–3 and 2×105 particles cm–3 at the stack emission from a municipal waste incinerator (MWI) with waste capacity of 7.5×103 kg h–1 and low heat value of 10.9 MJ kg–1. The flue gas treatment included selective non-catalytic reduction, an electrostatic precipitator, a spray absorber system and a fabric filter. About 99% and 65% of particles with size below 2.5 μm were due, respectively, to submicron particles and to ultrafine particles with a mode in between 80–90 nm. The latter, in terms of mass, represented less than 5% of PM2.5. Another study [25] reported particle size distributions and total concentrations before the fabric filter and at the stack of an MWI with waste capacity of 10–12×103 kg h–1 and low heat value greater than 15 MJ kg–1. Maximum PN concentrations of 2.7×107 particles cm–3 and 2.0×103 particles cm–3 and size number distributions with modes at about 150 nm and 90 nm were measured before and after the fabric filter, respectively. The relative metal mass concentration in ultrafine particles varied according to particle sizes. The contribution of elements with lower boiling point temperatures (below 1200°C) such as As (10.8–21.4%), Cd (5.3–10.9%), V (3.9–7.7%), and Zn (9.5–23.6%) decreased with increasing particle size (50, 100, 150, and 200 nm) whereas the contribution of elements with higher boiling point temperature (above 1200°C) such as Co (2.9–6.5%), Cr (6.6–11.1%), Fe (13.5–18.3%), Sb (1.9–5.2%), Cs (2.2–5.6%), Eu (2.4–3.4%), Sc (1.8–5.6%), Sm (1.5–4.5%), Th (0.9–2.1%), and Yb (1.2–3.8%) increased with particle size. Hg (0.9–1.0%) and Ni (0.5–0.6%) were absent in 50 nm particles and were present at about the same % concentrations in 100, 150, and 200 nm-particles.
Transport emissions
Ships
Corbett et al. [26] estimated that PM ship emissions are responsible at a global scale for approximately 60,000 cardiopulmonary and lung cancer deaths annually.
Emissions from ship diesel engines mainly consist of organic carbon, ash, sulphate, and metals such as V, Ni, Al, Ca, and Fe [27, 28]. Metals are present as free metals, carbonates, oxides hydrates, or labile complex with organic ligands [29].
Particle emission factors, expressed as number of particles emitted per kg of burnt fuel, range from about 8.8±1.0×1015 kg–1 to 6.2×1016 kg–1 [30–34].
Healy et al. [35] studied ship emissions at distances of 150–600 m from shipping berths in a port and measured PN concentrations as high as 2.1×105 particles cm–3 during a ship exhaust event.
Petzold et al. [36] reported PN concentrations in the exhaust gas of 1.26±0.51×1015 cm–3, and bimodal size number distributions with modes in the ultrafine size range at about 15 nm and 50 nm. Two main modes below 40 nm and in between 70 and 100 nm have also been reported by Isakson et al. [28].
Airports
Average UFP number concentrations of 3.1×104 particles cm–3 and 4.2×104 particles cm–3 have been measures at two terminals (concentration baseline of 1.0×104 particles cm–3 and 2.0×104 particles cm–3) of Heathrow Airport (UK), with peak levels as high as 7.0×104 particles cm–3. For two terminals of the Stockholm airport, average UFP concentrations of 5.1×104 particles cm–3 and 3.6×104 particles cm–3 have been reported with peak values above 3.0×105 particles cm–3. Steeply increasing values were observed during take-off events, returning to original values after about 20 s. UFP concentrations measured at a blast fence in Los Angeles international airport increased by a more than 100-fold factor during take-off with spike values above 1.0×107 particles cm–3. 30 nm particle concentrations have been shown to increase from below 5×103 particles cm–3 during the blast fence approaching phase to above 1.5×104 particles cm–3 during take-off in a 3 s time span [37].
In an U.S. EPA study, PN size distributions measured at a distance of 30 m behind aircraft engine exits were monomodal, with modes ranging from about 3 nm to about 100 nm, respectively. Geometric mean diameters (GMD), ranged from 10 to 35 nm, and closely followed the engine fuel flow rate variation. The largest GMDs were obtained at high power conditions; GMDs of about 10–20 nm and up to 35 nm were measured at lower (<2000 kg h–1) and higher fuel flow rates (6000–7000 kg h–1), respectively [38, 39].
Black carbon concentrations have been reported to vary in a well correlated way with the engine power thrust, with peak values ranging about from 0.3 to 7.0 mg m–3 [38].
PAH particle bound concentrations have been shown to follow the changes of engine power with peak values from 200 to 1,000 ng cm–3 [38].
Depending on the fuel type and flow rate and on the average rated power, average emission indexes for PN, particle mass, black carbon, and PAHs have been estimated in the range, respectively, from 1×1015 to 1×1017 particle kg-1 fuel burned, from 10 to 550 mg kg–1 fuel, from 62.5 to 592 mg kg–1 fuel, and from 0.00853 to 3.25 mg kg–1 fuel [38].
The total elemental emissions index (including Mg, Si, P, S, Cl, K, Ca, Ti, Cr, Mn, Fe, Ni, Cu, Zn, Br, Ag, In, Sb, Te, I, Tl) range from 5.3 to 27.5 mg kg fuel–1, amounting to about 2–7% w/w of total PM. 54–80% of the elemental mass was represented by sulphur [38, 39].
Hudda et al. [40, 41] clearly addressed the impact of airport emissions on outdoor air quality in Los Angeles and in Boston. They reported a 2-fold increase over background values of PN concentrations in a 60 km2 area extending 16 km downwind and a 4- to 5-fold increase at 8–10 km downwind the Los Angeles International Airport. At 8 km downwind, the authors reported concentrations above 7.5×104 particles cm–3, more than the average freeway PN concentration in Los Angeles.
In Boston, they measured median PN concentrations ranging from 8.2×103 to 3.7×104 particles cm–3 and from 7.2×103 to 1.6×104 at sites 3.7–10.0 km far from the Logan International Airport, respectively, in conditions of downwind advection of airport-related emissions and in unaffected conditions.
Habre et al. [42] in a randomized cross-over study involving walking activity of adult volunteers in two Los Angeles parks inside the zone studied by Hudda et al. [40] (PN concentrations 2- to 4- fold higher than background values) and outside (control site), reported increased acute systemic inflammation following exposure to airport-related UFPs.
Road transport
In urban environments, UFPs represents about from 70% to above 90% of total PN concentration [4, 43–47].
According to the European Union [48], UFP emissions from transport sources contributed to 34% of the total EU27 PM0.1 emissions in 2008, by far more than those due to industrial combustion (12%) and to power generation processes (4%).
The 2010 PN emissions due to road-traffic in EU28 have been estimated to range, in a well correlated way with the population (R2 = 0.91), from 6.24×1023 yr–1 for Malta to 1.35×1026 yr–1 for France. In a megacity such as Delhi, the PN emissions were about 32-fold and 11-fold higher than those due to Malta and France, respectively [49].
IARC [50] has considered sufficient the evidence in humans for the carcinogenicity of outdoor pollution and has classified both outdoor pollution and PM in outdoor pollution as carcinogenic to humans (Group 1).
UFP concentrations and size distributions quickly change with time and with the distance from the pollution sources [4, 51]. Such variability depends on several variables such as the types of fuels, engines, and lube oils, the vehicle maintenance degree, the ambient temperatures, the engine thermal history, and the driving phase. Near to traffic nucleation particle concentrations increase very fast within a few seconds and decrease within tens of second. The time scale of such variation is determined by the processes of dilution, nucleation of semivolatile organic compounds (SVOCs) to yield new particles, SVOC condensation and particle coagulation. As a result of that, the exposure pattern, near to traffic, may be represented as a sequence of short-term peak exposures.
Avino et al. [52] reported that the UFP concentrations correlate well with the traffic flow intensity and with other pollutants such as PAHs and nitrogen oxide, in downtown Rome. UFP level increases during the cold season, reaching for few seconds concentrations above 105 particles cm–3 in proximity of traffic and in relation to car exhaust plume emissions. The authors reported a marked seasonal pattern with higher values in the cold seasons and the clear presence of two concentration peaks at the heaviest traffic-hours. The second of them is less pronounced in warmer months than in cold period, possibly due to winter temperature inversions leading to low boundary layer height.
Gualtieri et al. [53] in the framework of the “Carbonaceous Aerosol in Rome and Environs (CARE)” experiment [54], exposed in field air liquid interface cultured BEAS-2B cells (representative of the distal bronchioles) to environmental concentration of PM. They observed significant association of anti-oxidant genes with secondary and aged aerosol and cytochrome activation with primary and PAHs enriched ultrafine particles.
Domestic heating
Residential heating is a well-known source of outdoor air pollution, even if the emissions differ widely according to the type of fuel and the combustion technologies. In particular, solid fuels such as coal, wood, and other biomasses are recognized as major causes of global PM pollution heating-related [55]. Recently, indeed, the growing levels of environmental emission of PM has been related to the increase of biomass consumption in Europe due to both economic and ecological reasons [56]. The impact of residential biomass combustions on air quality was estimated until to 40% of the total burden of airborne particle mass levels [57].
PM pollution related to residential heating significantly differs also from an area to another. Over the world, it has been estimated that residential heating stoves and boilers releases less than 10% of outdoor PM2.5, with the main contribution from biomass and household coal burning heating [55]. In specific areas, where wood combustion is the major source of ambient particulate, higher PM levels are recovered; for example, a study performed in Italy demonstrated that residential heating caused 18–76% of PM10 in urban areas and until to 85% in rural areas [58].
Residential heating is also linked with UFPs emissions, as evidenced by the great number of studies in field performed in the last decade. Based on the estimation of the PN emission, that is a useful way for assessing the contribution of different sources; domestic combustion is the third source of total PN emission, preceded by road traffic and non-road traffic contributions [49]. Also, for this specific PM fraction, a relevant source is the combustion of solid fuels, such as coal [59], wood [60], and other biomasses [61]. With the respect to biomass fuels, several studies demonstrated that both household-scale and larger district heating facilities emit relevant amounts of UFPs [62]. On a quantitative point of view, the UFPs contribution to the total PM emission relatively increases with the improvement of the operational practice efficiency of the facility [63].
UFPs generated by solid fuel combustion differ in chemical composition due to the kind of fuel. In particular, UFPs derived from biomass combustion are composed mainly of sulfate, nitrate, chloride, sodium, potassium, calcium, magnesium, ammonium, zinc, elemental carbon, and particulate organic matter [64, 65]. Coal-derived UFPs contain a large number of elements such as Na, K, Mg, Ca, Ti, Mn, Fe, Co, Ni, Zn, V, Cr, Cu, Sb, As, Se, S, Cl, and carbon [66].
INDOOR SOURCES
Residential environments
In the absence of indoor aerosol sources, indoor UFP concentrations are determined by the penetration of outdoor particles. The infiltration Factor (F in ), defined as the equilibrium fraction of outdoor particles that penetrate indoors and remain suspended, is expressed as [67]:
Actually, indoor aerosol derives both from the permeation of outdoor particles and from the emissions of indoor sources [70, 71]. As to the latter, there is a wide range of combustion and non-combustion sources that may be active in residential environments, sometimes causing PN concentration to reach levels well above than those encountered in urban environments. Among the first are mosquito coil, incense, and citronella candle burning, tobacco cigarette smoke, and cooking activities; among the second, domestic appliances operated by brush electric motors, hot flat irons, spray air fresheners, and heat not burn smoking devices [72–80]. Considerably higher aerosol emissions occur from combustion sources than from non-combustion ones. Average UFP increments over the background value are highest for electric appliances (from 5% to 12%) and lowest for combustion sources (as low as –24% for tobacco cigarette smoke). On the contrary, average increments in Alveolar Deposited Surface Area (ADSA) are highest for combustion sources (as high as 3.2×103 μm2 cm–3 for meat grilling without exhaust ventilation) and lowest for electric appliances (20–90 μm2 cm–3). This is due to both the high PN concentrations reached when combustion sources are active and to the relevant particle size distributions that contain an important fraction of larger particles (>100 nm) [77]. ADSA is a health relevant parameter because UFPs, compared with larger-sized particles of the same chemical composition, can generate higher level of ROS, due to their higher surface area per unit mass and to their surface reactivity [2].
Buonanno et al. [81] measured PN emission factor for unit mass (NEF m ) ranging from 5.2×1010 to 5.6×1010 particles min–1 g–1 and from 8.5×1010 to 9.6×1010 particles min–1 g–1, respectively for fat-rich food grilling and frying. For vegetable foods, the authors reported NEF m ranging from 3.6×1010 to 3.9×1010 particles min–1 g–1 and from 7.2×1010 to 7.7×1010 particles min–1 g–1, respectively for vegetable food grilling and frying. With the mechanical ventilation in operation, the authors measured for fat-rich foods maximum number concentrations, for 100 g of food, ranging from 1.8×105 to 1.9×105 particles cm–3 and from 2.7×105 to 2.8×105 particles cm–3, respectively for grilling and frying. For vegetable foods, they measured from 1.5×105 to 1.6×105 particles cm–3 and from 2.3×105 to 2.4×105 particles cm–3, respectively for grilling and frying.
Manigrasso et al. [77], for 150 g meat grilling, measured monomodal size number distributions with a mode at about 40 nm, superimposed to the one at about 10 nm, due to the emission of gas burning [82]. With mechanical ventilation on, they measured PN concentrations as high as 2.6×106 particles cm–3 and as high as 4.1×107 particles cm–3 after a second meat grilling session without mechanical ventilation. To put these figures into perspective, it is worth observing that in proximity of traffic, where aerosol in car exhaust plumes evolves on the time-scale of few seconds, lower peak PN concentrations (UFPs contribution above 90%) as high as 4.0×105 particles cm–3 (1 s time resolution measurements) have been reported [52].
During the operation of appliances operated by brush electric motors, electric arc discharge occurs between the copper windings in the rotor and the graphite electrodes (brushes) in the stator. As a consequence, aerosol with mode at about 10 nm is emitted through the surface vaporization of the copper windings [77, 84]. Particles size distributions with slightly lower modes (6–9 nm) have been measured by Wallace et al. [85]. Particles with greater sizes (e.g., 15–30 nm) probably derives from particle coagulation and/or contribution from the graphite brushes [86]. Spike PN concentrations and ADSA as high as 106 particles cm–3 and 200 μm2 cm–3 and as high as 1.1×105 particles cm–3 and 110 μm2 cm–3, were measured (1 s time resolution) on switching on, respectively, an electric drill and a vacuum cleaner [77].
Following the operation of this kind of motors, Szymczak et al. [84], by means of MOUDI impactor samplings, measured broad copper size mass distributions, with modes of about 1-2 μm, extending down to the UFP size range. Coherently, Manigrasso et al. [87], through high-resolution field emission scanning electron microscopy, observed that copper nanoparticles were present both as single NPs (20–40 nm) and aggregated in clusters in the μm sizes range.
In addition to the sources presented above, laser printer devices (LPDs) have been addressed by many studies as sources of UFP emissions in indoor environments. He et al. [88], based on a 150-page print job at 5% toner coverage, measured particle number emission rates from 30 LPDs (21–52 page min–1printing speed) varying from 4.25×109 to 3.30×1012 particles min–1. Similar emission rates have been reported by Scungio et al. [89] (3.39×108 particles min–1–1.61×1012 particles min–1) who also measured surface area and mass emission factors, respectively, of 1.06×100–1.46×103 mm2 min–1 and 1.32×10–1–1.23×102 μg min–1. The authors reported particle number size distributions with modes of the emitted particles ranging from 10 to 60 nm with a median value of 34 nm. In agreement with Scungio et al. [89], Wensing et al. [90] in chamber experiments, measured bimodal number size distributions from laser printer emissions, with a smaller mode below 10 nm and a broader one extending from about 40 nm up to 100 nm. Moreover, the authors measured UFP emissions with similar size distributions, even when laser printers were operated without toner or paper. In this case, the first mode was stronger than in measurements with paper and toner. Based on that, the authors proposed that the high-temperature fuser unit (the device that fixes toner on the paper surface) was the main UFP source. They argued that UFP formation proceeds through nucleation and condensation of volatile organic compounds (VOCs) or SVOC released from the chassis (flame retardants, lubricants, or plasticizers) or emitted by the fuser (siloxanes or fluorinated compounds). Indeed, among the LPD emissions, the authors detected tri-xylyl phosphate, a phosphororganic flame retardant used for the fire protection of plastic materials, in amounts correlated with the number of single-page print jobs. Coherently, He et al. [88] reported average outside fuser temperatures ranging from 130°C to 210°C and showed that particle number emissions during printing follow the fuser temperature variation.
Barthel et al. [91] reported the presence in the emissions of LPDs also of a fraction non-volatile at 400°C, amounting from 0.2% to 1.9% of the total number of emitted particles at room temperature and containing Si, S, Cl, Ca, Ti, Cr, and Fe as well as traces of Ni and Zn in size fractions from 30 to 160 nm. Moreover, the authors measured also significant Br concentrations deriving from the brominated flame retardants contained in plastic housings of the fuser units.
Martin et al. [92] measured at eight commercial photocopy centers, daily geometric mean UFP concentrations varying from 3.7×103 to 3.4×104 particles cm–3, up to 12 times greater than the relevant background values, with transient peaks above 1.4×106 particles cm–3, and weekly size number distributions with count median diameters ranging from 28 to 38 nm. Sampled UFPs contained 6–63% organic carbon,<1% elemental carbon, and 2–8% metals, including Fe, Zn, Ti, Cr, Ni, and Mn.
The recent technology of 3D printing allows to manufacture three-dimensional items, by extruding through a heated nozzle a thermoplastic polymer and depositing it, layer upon layer, on a moving bed. In this process, UFPs are emitted. Stabile et al. [93] investigated ten different thermoplastic materials, extruded at different temperatures. They reported UFP emissions with modes in the 10–30 nm range for all the materials and emission rates up to 1×1012 particles min–1, increasing with increasing of the extrusion temperature (180–240°C). The authors also reported ADSA doses as high as 200 mm2, for a 40 min long 3D printing task (16 mm s–1 printing speed). Azimi et al. [94] reported UFP emission rates (108–1012 particles min–1) in agreement with Stabile et al. [93] and VOC emission rates of 2–180 μg min–1 for caprolactam (from nylon-based and imitation wood and brick filaments), 10–110 μg min–1 for styrene (from acrylonitrile-butadiene-styrene and high-impact polystyrene filaments), and 4–5 μg min–1 for lactide (from polylactic acid filaments).
Occupational settings
Nanomaterials and nanoparticles
About 1,800 products containing NMs were worldwide commercialized in 2014 [95]. In about 49% of them, information on the NMs used was lacking. Metals and metal oxides were contained in 37% of them. About 5% were the products containing silicon or silica. About 5% of the commercialized products were formulated with carbonaceous NMs, such as carbon black and single- or multiwalled carbon nanotubes. On mass basis, titanium dioxide (TiO2), silicon dioxide, and zinc oxide were the most worldwide produced NMs, whereas silver NPs were the most widely diffused, being contained in 24% of the commercialized products.
Among metal and metal oxide category, TiO2, due to its photolytic properties, is extensively used in solar cells, paints, and coatings. TiO2, as well as zinc oxide (ZnO), is used also in cosmetics and sunscreens [96]. Cerium dioxide is widely used in many sectors including catalysis, fuel cells, phosphor/luminescence, abrasives for chemical mechanical planarizations. These applications rely on the high thermodynamic affinity for oxygen and sulfur, its potential redox chemistry involving Ce(III)/Ce(IV), and the unique useful absorption/excitation energy bands associated with its electronic structure [97].
Gold NPs, due to their ability of being functionalized with thiol groups, may be conjugated with a variety of biological molecules. Therefore, they are used for diagnostic and biomedical applications [98, 99].
Quantum dots (QDs) are semiconductor nanocrystals [100]. QDs made from cadmium selenide, cadmium telluride, indium phosphide, and zinc selenide, thanks to their optical and electrical properties, are used in medical bioimaging, targeted therapeutics, solar cells, photonics and telecommunication [101–103].
Single-walled carbon nanotubes and multi-walled carbon nanotubes, due to their mechanical properties, thermal conductivity, electrical conductivity, and high aspect-ratio (length to diameter ratio), are employed in a variety of products such as aircraft and automotive components, electronic components, supercapacitors, battery and fuel electrodes, catalyst supports, and air and water filtration [96, 104].
Human exposure to NMs not only occurs via inhalation, but also through dermal contact of consumer products [105] and via food ingestion [106]. Specifically, Wang et al. [107] in an acute oral toxicity study on mice, showed that TiO2, after uptake by the gastrointestinal tract, is mainly retained in the liver, spleen, kidneys, and lung tissues. Due to the widespread use of TiO2 in cosmetics, its dermal absorption as well has been studied by a number of studies [108–110].
Paralleling with the diffusion of such a wide range of NMs, concerns have been rising on their health effects [2, 111]. In this regard, IARC has classified TiO2 as possibly carcinogenic to humans (group 2B) based on inadequate evidence in humans and sufficient evidence in experimental animals, due to the increased incidence of lung tumors in intratracheally- and inhalation-exposed rats [112]. Czajka et al. reviewed the studies on the effects of TiO2 on the central nervous system [113]. The authors concluded that TiO2 NPs may accumulate in the brain, causing brain damage and neurotoxicity. They observed that TiO2 NPs triggers different signaling pathways that lead to the apoptotic processes through mitochondria-dependent pathway. Besides, microglial activation occurs, with consequent release of proinflammatory cytokines, further neurodegeneration and brain injury.
Lin et al. [97], after exposing human lung cancer cells to cerium dioxide NPs, observed a dose-dependent and time-dependent decrease of the cell viability significantly.
Hardman [103] addressed that QD toxicity, absorption, distribution, metabolism, and excretion depend on a range of factors such as their shape, shell and core composition, outer coating bioactivity, and size. The author pointed out that the inhalation route of exposure, occurring in occupational environments, may be harmful, considering that QDs may be incorporated into cell via endocytosis. For the same reason, he argues that dermal and ingestion exposures as well should be carefully considered, although the relevant risks are not yet documented. Moreover, QDs may be systemically distributed and may accumulate in organs and tissues. Tsoi et al. [114] pointed out that QDs may be apparently toxic in vitro but safe in vivo, due to the different NP-cell interaction in cell culture conditions rather than in vivo. Different doses may reach the target cells in vivo compared with in vitro; moreover, surface chemistry and size modifications may occur as NPs pass through the body [114]. Hoshino et al. addressed that QD surface coating may be detached under acidic and oxidative conditions in endosomes and released into cytoplasm, playing a role in the NP toxicity [103, 115].
Carbon nanotubes (CNTs) toxicity has been put in relation with some physicochemical characteristics such as their length diameter, surface area, tendency to agglomerate dispersibility in solution, presence and nature of catalyst residues and chemical functionalization [116–121]. The CNT toxicity has been attributed to the production of oxygen radicals [122]. Cellular apoptosis induced by CNTs has been reported to proceed through lysosomal and mitochondrial damage. The lysosomal digestive enzyme released finally damage the entire cell. CNTs may decrease the mitochondrial membrane potential, resulting in the formation of ROS. A high level of ROS may in turn damage cells by altering protein structure, disrupting DNA, interfering with signaling functions, and modulating gene transcription. This can eventually result in cancer, renal diseases, neurodegeneration, cardiovascular or pulmonary diseases [123].
Ultrafine particles
UFPs are generated in occupational settings following high temperature and mechanical treatments of materials, due to unintentional heating of semi-volatile materials, and due to combustion processes. Notwithstanding their diffusion, neither number nor surface area UFP exposure limits currently exists.
As to high temperature treatments, Elihn et al. [124] at a site a with background particle concentrations of about 7,000 cm–3, during paving operations, measured particle levels as high as 2.2×105 cm–3, with an average value of 3.4×104 cm–3 and particle size number distributions with a geometric mean diameter of about 70 nm.
To exclude the contribution due to traffic exhausts, the authors carried out measurements at an asphalt plant and found about the same peak concentrations values as at the paving site. They pointed out that UFPs derived to greatest extent from the asphalt fumes due to the high temperature (about 150°C), because as the paving was interrupted, the particle concentration decreased.
Zhang et al. [125] studied the exposure to UFPs arising from gas metal arc welding in an automotive plant and found NP number concentration, lung-deposited surface area concentration and mass concentration significantly higher than the relevant values, before welding. At a welding booth equipped with a local exhaust hood (in addition to an industrial fan), they measured average PN and lung deposited surface area concentrations, respectively, of 1.8×105 particles cm–3 (background value of 4.5×104 particles cm–3) and 440.04 μm2 cm-3 (background value of 106.78 μm2 cm–3). The authors measured unimodal particle size number distribution, with mode at 190 nm and found that UFPs accounted for about 61% of the total number concentrations measured (3 min scan-time measurements). Welding particles were mainly present as chain-like agglomerates of primary particles formed from vaporization and vapor phase nucleation of base metal and of the wire electrodes. They consisted primarily of Fe, Mn, Zn, and Cu.
Avino et al. [126, 127] carried out aerosol number size distributions measurements with 1-s time resolution during metal inert gas (MIG) welding and pointed out the relevance of nucleation mode particles. Such particles before welding represented about 7% of submicrometric particles; after about 40 s from the welding start, the percent contribution of nucleation mode particles (measured as total concentration from 6 to 16 nm) increased to 60% and their concentration passed from about 1.6×103 to 1.0×106 particles cm–3.
Buonanno et al. [128] measured, at about 3 m from the source operation, at an automotive plant without general ventilation system, average PN concentrations as high as 9.6×104, 1.9×105, 2.3×105 particles cm–3 and lung deposited surface areas as high as 4.4×102, 6.2×102, 3.2×103 μm2 cm–3, respectively for gas metal arc welding (GMAW), resistance welding and oxyacetylene welding. Local exhaust ventilation was absent with the exception of the 4.4×102 μm2 min–1 measurement for GMAW, for which the author reported limited exhaust ventilation. The authors estimated overall automotive plant emission factors of 2.8×1015 particles min–1 and 7.0×106 μm2 min–1 for PN and surface area concentrations, respectively. They measured with 1 s time resolution unimodal size number distributions for resistance welding, with mode at 60 nm and bimodal size number distributions for GMAW, with modes at 10 and 93 nm.
Graczyk et al. [129] measured at the breathing zone of apprentice welders performing Tungsten inert gas welding a mean PN concentration of 1.69×106 particles cm–3, with a mean geometric mean diameter of 45 nm. They found on average that 92% of the particle counts were below 100 nm. The authors reported primary welding particles of spherical shape in the nanoscale range (about 20–50 nm), present also as chainlike agglomerates with fractal geometry and observed elevated concentrations of tungsten, very likely due to electrode consumption.
Evans et al. [130] measured, at an automotive grey iron foundry, UFPs number concentrations that varied from 1.9×104 to 3.5×106 particles cm–3, in relation with the batch nature of particle-generating activities such as melting and pouring and with the variability in process emissions and frequent work interruptions.
Kim et al. [131] performed aerosol measurements at a rubber manufacturing factory, where the production cycle involved feeding waste tire into a mold, curing by pressing and heating at 150°C, demolding, and a final operation of drilling holes and cutting ends of the items produced. The authors reported that workers were exposed at high concentrations of rubber fumes starting from demolding operation to the end of the last task. At this stage, they reported 15 min-peak concentrations of alveolar deposited surface area and total PN, as high as 640 μm2 cm–3 and 5.4×105 particles cm–3, with a count median diameter of 26 nm, and with a 95% contribution due to UFPs.
Vosburgh et al. [132] carried out aerosol measurements in the breathing zone of the sealing workers of a facility manufacturing waterproof polytetrafluoroethylene apparels. They reported PN concentrations ranging from 1.5×105 to 7.8×105 particles cm–3, considerably higher than those measured in the office area (1.2×104 particles cm–3) and a number median particle diameter of 25 nm, measured at the final discharge of the facility’s ventilation system.
Karlsson et al. [133] measured the PN concentrations and size distributions in car repair shops during the grinding, welding, and cutting operations. They reported particle concentrations as high as 107 particles cm–3, with geometric mean diameters of about 50 nm. The author found low volatility isocyanates, mainly in the particle phase, whereas relatively high volatility isocyanates were present both in the particle and in the phases.
Voliotis et al. [134], during the firing (up to above 600°C) of unpainted and unglazed ceramics, measured average and peak PN concentrations, respectively, of 1.6×105 and 6.5×105 particles cm–3 (average background concentration of 2.0×104 particles cm–3), with mean size ranging from 30 to 70 nm. During the firing of painted and glazed ceramics, the authors measured average and peak PN concentrations of 2.5×105 and 1.6×106 particles cm–3, with mean size ranging from 15 to 40 nm. In both the firing operation, Energy-Dispersive X-ray spectra revealed the presence of Si bearing particles and the constant presence of lead in the particles collected during glaze firing.
Manufacturing operations involving mechanical energy may represent emission sources of UFPs due either to heat build-up and possible vaporization of semi-volatile materials or to unintentional direct aerosol dispersion from materials being processed. Unintentional heating of semi-volatile materials as well may represent an UFP emission source. As to the possible vaporization of semi-volatile materials, an example is represented by metal working fluids (MWFs). MWFs are semi-volatile and vaporize when they are heated, as a result of processes involving mechanical energy (e.g., machining, impaction, and compression) or thermal treatments (e.g., quenching and annealing). In these cases, UFP generation occur through a mechanism of evaporation and condensation [135].
Heitbrink et al. [136] reported UFP emissions from machining operation due to MWF escaping from poorly fitting enclosures, in an engine machining facility. The authors measured PN concentrations in the range from 1.2×105 to 2.2×105 particles cm–3, with modes in between 23 nm and 100 nm. They argued that such emissions may possibly derive from thermal degradation of MWF during the heat- treatment of induction-hardening. They may also derive from MWF volatilization at the interface between the grinding tool and the item and subsequent condensation.
Wang et al. [137] examined the processes of forming, threading, and heat treatment involving the use of MWFs in the fastener manufacturing process. They measured average PN concentrations respectively of 2.1×105, 1.4×105, 3.5×105 particles cm–3 with count median diameter of 26.9, 23.2, and 22.5 nm, in comparison with 1.3×104 particles cm–3 and 41.1 nm measured in the background ambient environment.
As to the unintentional direct aerosol dispersion, an example is represented by dental composites. Dental composites for tooth restoration are made of a polymer and inorganic filler particles and contain considerable amounts of nanoscale filler particles, such as silica nanoparticles, TiO2 nanoparticles, methacrylate-modified silicon-dioxide-containing nanofiller, nanoscaled and highly aligned fibrillar silicate single crystals into nylon 6 nanofibers, to name a few [138, 139]. Van Landuyt et al. [140] pointed out that both dental personnel and patients may inhale nanosized dust particles due to abrasive procedures carried out on dental composites. During grinding procedures, they measured PN concentrations ranging from 5×106 to 2×107particles cm–3, consisting of filler particles or resin or both, with median diameters between 38 and 70 nm, UFP content ranging from 70% to above 95%.
During abrasive treatment of carbon and glass fiber-reinforced epoxy-composites, Jensen et al. [141] measured average PN concentrations of 3.9×104 cm–3 and 1.7×106 particles cm–3, with particles primarily below 100 nm size range, respectively, outside and inside a tent enclosing a grinding station.
Mølgaard et al. [142] during the spraying of mold-release agent in a workshop performing polyurethane molding, measured PN concentration burst of about 105 particles cm–3 (outdoor particle concentrations were below 104 particles cm–3) with a count median diameter of about 10 nm. The authors argued that most likely, spray droplets, hitting a heating surface, evaporated and then new particles nucleated from the vapor phase. Alternatively, they hypothesized that nucleating compounds were formed via chemical reactions.
As to combustion processes, exposure to diesel exhaust (DE) is of particular health relevance, because IARC [143] has classified DE as carcinogenic to humans (Group 1). Lewtas and Silverman [144] reported that about 1.4 million workers in the United State and 3 million workers in the European Union are occupationally exposed to DE. Several occupational groups are exposed to DE, such as miners, professional drivers, railroad workers, vehicle mechanics, heavy equipment operators, dockworkers, tunnel workers, firefighters, farmers, and shipping engineers [144, 145].
Bujak-Pietrek et al. [146] carried out aerosol measurements in the service room of a bus depots. The authors reported mean PN concentrations of 7.6×103 and 1.3×105 particles cm–3, respectively, before and during bus servicing procedures. During service, mean deposited particle surface area concentration were 95.97 and 356.46 μm2 cm–3, respectively, for the tracheobronchial and the alveolar regions. Wheatley and Sadhra [147], in depots where diesel powered fork-lift trucks where used, measured background time weighed average (7–10 h) UFP concentrations ranging from 5.0×104 to 2.2×105 particles cm–3, with peak concentrations above 5.0×105 particles cm–3 (1 min integration time).
Debia et al [148] evaluated the occupational exposures of port-gate controllers to DE from container trucks. They reported average daily PN concentrations ranging from 1.6×104 particles cm-3 to 6.7×104 particles cm–3 with peak concentrations as high as1.0×107 particles cm–3 due to a truck with an old exhaust system. More than 99% of particles measured were UFPs, with the main particle size fraction being between 20–40 nm.
Fruin et al. [149] measured diesel PM inside of vehicles and reported that the average concentration was from 4 to 5 times the ambient concentrations throughout California. Coherently, Mayer et al. [150] addressed that filters currently mounted in car ventilation systems are usually inefficient for removing very small particles below 0.5 μm and showed that PN concentration inside a vehicle is almost the same as outside.
Fruin et al. [151] pointed out that that from 33% to 45% of Los Angeles population exposure to UFP occur during the 6% of time people spend in vehicles. Therefore, it can be argued that professional drivers are to greater extent exposed.
As to other environmental settings where combustion aerosol, other than DE, maybe encountered, Jordakieva et al. [152] measured PN concentrations, at police shooting sites, ranging from 3.3×105 particles cm–3 to 7.6×105 particles cm–3, with monodispersed size distributions and mode ranging from 56 to 101 nm. In the inhalable aerosol fraction collected the authors detected Pb, Ba, Cu, and Sb.
Baxter et al. [153], during overhaul of simulated fire scenarios, measured PN concentrations ranging from 1.9×104 particles cm–3 (automobile passenger compartment fire) to 2.1×106 particles cm–3 (residential building-bedroom fire), with UFPs contribution ranging from 70% to 91%. The authors addressed that UFP exposure during fire suppression should be considered a potential contributing factor for coronary heart disease in firefighters, particularly during overhaul, where firefighters frequently remove respiratory protection.
Evans et al. [154] reported that although vehicle fires are suppressed quickly (<10 min), firefighters may be exposed to short duration, high particle concentration episodes during fire suppression and measured a maximum transient particle concentration as high as 1.21×107 particles cm–3, with a preponderant contribution due to UFPs.
Møller et al. [155] assessed the personal exposure to UFP of different occupational groups working at an airport. They reported significant differences among occupational groups of airport employee. Baggage handlers were exposed to 7 times higher average concentrations (geometric mean UFP concentration of 3.7×104) than employees mainly working indoors (geometric mean UFP concentration of 5×103 particle cm–3). Catering drivers, cleaning staff, and airside security were exposed to intermediate concentrations (geometric mean UFP concentration from 1.2 to 2.0×104 UFP cm–3).
Heitbrink et al. [136] reported that the geometric mean PN concentration was considerably higher (3.1×106 particles cm–3) when natural gas heaters were switched on than they were off (1.3×105 particles cm–3), in an Engine Machining and Assembly Facility.
HEALTH RELEVANCE OF PARTICLE SIZE
The causative link of urban aerosol pollution with respiratory and cardiovascular diseases has been well established by several studies [156–159]. Moreover, it is from about two decades that evidences have begun accumulating on the association of air pollution and adverse effects on the central nervous system [160].
The first evidence comes from the study of Kilburn [161], who reported deteriorated visual performance, impaired cognitive performance, and balance impairment among railroad workers and electricians exposed to DE in confined environments.
Later, other authors showed an association between living near major roadways and higher incidence of dementia [162], reported adverse neurological effects of traffic air pollution on experimental animals [163, 164], and demonstrated that DE particles in vitro can activate microglia and induce oxidative stress and neuroinflammation [165, 166].
Calderón–Garcidueñas et al. [167–169] have suggested that air pollution may play a role in the etiology of neurodegenerative diseases such as Alzheimer’s disease.
Costa et al. [163, 164] have pointed out that biochemical alterations associated with neurodevelopmental and neurodegenerative diseases may occur even after short-term exposures, so that occupational exposures as well are of concern. Indeed, neurological effects have been reported among welders exposed to welding fumes containing Mn [170, 171]. Mn accumulation has been reported in globus pallidi brain region and on the olfactory bulb of welders by Uchino et al. and by Sen et al. [172, 173]. Furthermore, Antonini et al. [170] addressed that manganese may be present in different oxidation states and have different solubility properties, consequently affecting the biological responses after the inhalation of welding fumes. The role of redox-active metals in neurodegenerative diseases, due to their ability to produce ROS, has been established by several studies for manganese [174–176], iron [177, 178], and copper [179–182].
In particular, hot spot Cu and Zn, at higher concentrations than Ca and Fe, have been observed, with high spatial correlation, in the amyloid-β plaques present in Alzheimer’s disease brain tissue samples [183]. Within this context, the size of inhaled particles is of primary importance, because it is the main parameter determining the efficiency of particle deposition in the regions of respiratory system, with consequences for the potential effects induced as well as for their disposition to extrapulmonary organs [2]. Furthermore, inflammatory effects have been reported to be inversely related to particle size [184–186].
As to the relevance to neurodegenerative diseases, several studies demonstrated the translocation to brain through olfactory bulb and olfactory nerve of intranasally instilled [2] and inhaled nanoparticles in experiment animals [187, 188]. Only recently, Maher et al [189] reported the presence in the human brain of combustion-derived magnetite and of other metal-bearing nanoparticles (sizes below 200 nm) with median shortest and longest diameters of approximately 14 and 18 nm, respectively. The authors hypothesized that such particles may bypass the typical way of uptake (circulatory system, blood-brain barrier) and reach the brain directly through the olfactory bulb.
Therefore, based on all these evidences, it is important to estimate the doses of NPs that can potentially deposit on the olfactory bulb, because they are candidate to possibly translocate to the brain. With this aim, the histogram reported in Fig. 1 represents the average values of the frequencies of the magnetite NPs retrieved by Maher et al. [189] in brain tissue samples, as a function of the longest and of the shortest particle diameters. The curves in the same figure show the number size distribution of particle doses deposited on the olfactory bulb per 1 min time exposure, for different emission scenarios encountered in occupational settings (MIG welding with exhaust ventilation) [127], in indoor environments (meat grilling with exhaust ventilation, mosquito coil burning, electric drill operation) [80] and in an urban traffic environment in downtown Rome [4], calculated according the equations reported in Manigrasso et al. [87]. These curves represent a demonstration on the deposition of particles on the olfactory bulb in some real scenarios and, as such, they are susceptible of variation depending on the relevant influential parameters.
Number size distribution of particle doses deposited on the olfactory bulb per 1 min time exposure for MIG welding with exhaust ventilation, meat grilling with exhaust ventilation, mosquito coil burning, electric drill operation (switching on/off once), compared with the average value of the frequencies of the magnetite NPs retrieved by Maher et al. [189] in brain tissue samples, as a function of the longest and of the shortest particle diameters (histogram).
Upon inhalation, about 3.3×107, 1.8×107, 1.4×106, 1.7×106, and 9.8×105 particles are respectively deposited on the olfactory bulb per minute of source operation. All the curves overlap well to the histogram, suggesting that UFPs released by emission sources present in indoor (both residential and occupational) and outdoor environments can possibly translocate to the brain. Of particular concern is the exposure to UFPs in residential environments, because people spend more than approximately 80% of their time indoors [190], and, such exposure is chronic, because it derives from routinely daily performed operations and involve the general population.
Moreover, chemical analyses performed throughout a year on 1-month averaged PM10 samples collected in private dwellings have shown that the appliances operated by brush electric motors cause indoor copper concentrations more than two–fold higher than in outdoor, with a predominant contribution of the Cu insoluble fraction [87]. Therefore, such particles, for the amount they provide to the insoluble fraction, may possibly undergo harmful tissue accumulation. This occurrence deserves particular attention in view of the role assigned to copper ions on the onset of Alzheimer’s disease [179–182].
CONCLUSION
Nanosized particles, be they unintentionally generated or purposely used in manufacturing processes, are ubiquitous in indoor and outdoor environments. They may be primary particles emitted directly by combustion processes or derive from condensation of semivolatile compounds or derive from processes where mechanical energy is involved. Their size allows them to deposit upon inhalation on the olfactory bulb from where they can possibly translocate to the brain.
NPs containing redox-active metals are of particular concern in view of their ability of generating ROS and promoting inflammatory processes, recognized to be at the base of neurodegenerative diseases. Among them, copper nanoparticles deserve particular attention due to the many causative associations reported between copper ions and Alzheimer’s disease. UFPs in residential indoor environments are of particular concern, because they derive from emission sources routinely active and involve the general population with a chronic exposure pattern.
DISCLOSURE STATEMENT
Authors’ disclosures available online (https://www.j-alz.com/manuscript-disclosures/18-1266r1).