Chapter 1: Occurrence of ENPs and Nanoplastics in Different Environmental Compartments: An Overview Free

One of the landmarks of the industrial development of the new millennium is related to nanoscience and nanotechnology. Thanks to their special properties, the use of nanomaterials (NMs) and nanoparticles (NPs) in industrial and consumer products is continuously increasing, with expectations of a global nanotechnology market that will exceed US$125 billion by 2024.¹  Indeed, engineered NPs (ENPs), which are defined as objects, intentionally produced and designed, with dimensions at the nanoscale (1–100 nm),²  are use in a wide range of applications, like agriculture,³  the military industry,⁴  the automotive industry,⁵  medicine,⁶  and cosmetics.⁷  One of the immediate consequences of this widespread use is that NPs have multiple pathways through which they can interact with different environmental compartments⁸  (i.e. air, water, soil and living organisms) during their whole cycle life. In this context, the identification of the occurrence of ENPs is essential for a reliable risk assessment, but at the same time is one the main environmental challenges nowadays, due to the low concentration of these emerging contaminants in environmental compartments, which make them invisible to the majority of analytical techniques.⁹ 

In addition to the most often used NPs (metallic, metal-based, quantum dots, organic) a new kind of environmental nanopollutant has attracted a great deal of attention over the last few years: nanoplastics. This term refers to plastic debris of mixed composition and shape, having colloidal behaviour, and with sizes ranging between 1 nm and 1 µm.¹⁰  Nanoplastics result from the degradation of larger plastic debris like microplastics (plastic particles below 5 mm),¹¹  and potentially present in all environmental compartments. Furthermore, their potential role as carriers of other pollutants may not be discarded.¹²  Their identification in environmental compartments is even more challenging than in the case of ENPs, due to their chemical nature.

Natural NPs (NNPs) are also widely present in nature, and the distinction between them and man-made NPs in environmental samples is not always possible. In addition, data regarding the background level of naturally occurring NPs is not available. However, the metal content ratio in some NNPs is known, which helps to distinguish them from ENPs when using inductively coupled plasma mass spectrometry (ICP-MS) based techniques.

This chapter provides a brief overview of the current knowledge on the presence of nanoparticles and nanoplastics in different environmental compartments like air, water, soil and living organisms. It is based on the literature data from peer reviewed journals and intends to highlight the most pressing problems and research needs related to these emerging environmental contaminants.

While the occurrence of incidental NPs in the atmospheric environment has been largely studied, interest in the potential presence of ENPs in air samples, especially in urban environments, has increased over the last few years.¹³ 

Bäuerlein et al. studied the presence of metal-based (Mo, Ag, Ce, W, Pd, Pt, Rh, Zn, Ti, Si, B, Fe and Cu) NPs in air samples collected in four different location in The Netherlands.¹⁴  The authors found that the majority of metals were in the size fraction larger than 100 nm. The presence of platinum group elements (PGE) was explained by their use as catalytic converters. Other studies have also demonstrated the occurrence of PGE elements in nanoparticulate form in particulate matter (PM).¹³  Baalousha et al. studied the presence of different NMs in samples collected during a hazy day and a clear day in Shanghai, China.¹³  FeO NPs with sizes ranging from 5 to 150 nm and PbS NPs with sizes ranging from 100 to 200 nm were found in the PM_2.5 samples, although their anthropogenic origin was not confirmed.

Carbon-based NPs have also been detected in air samples. For instance, C₆₀ and C₇₀ have been found in aerosols above the Mediterranean Sea with median concentrations of 0.06 ng m⁻³ and 0.48 ng cm⁻³, respectively.¹⁵  C₆₀ has also been detected in outdoor samples collected in different places in Vitoria-Gasteiz (Spain), with a maximum concentration of 2.27 pg m⁻³.¹⁶  The occurrence of functionalised fullerenes was investigated in The Netherlands.¹⁴  Only one fullerene (C₆₀) was found in one of the four locations studied, at a concentration (5 pg m⁻³) similar to those measured in exhaust from various fuels.¹⁷ 

Concerning plastics particles, some publications have shown the occurrence of airborne microplastics in urban areas, mainly polyethylene terephthalate (PET) fibres,^18,19  although a study performed in remote areas of the Pyrenees reported a higher amount of sheets and fragments rather than fibres.²⁰  More recently, the evidence of 0.09–0.66 microplastic particles per m³ was demonstrated in four summer months from the Pic du Midi Observatory (Pyrenees) at 2877 m above sea level, which is defined as a clean station.²¹  In the same sense, plastic nanoparticles have also been identified in high altitude environments. Materic et al. detected and identified PET nanoplastics in snow pits in the Austrian Alps by using thermal desorption-proton transfer reaction mass spectrometry (TD-PTR-MS).²²  The presence of nanoplastics at high altitude shows airborne transport of these environmental nanopollutants.

Several studies have focused on the analysis of the occurrence of NPs in surface waters. One of the first screening assessments of metal-containing NPs in a surface water was the study performed by Hethmar and Pergantis in an urban runoff (Las Vegas Wash).²³  Ag, Ti and Ce-bearing particles were detected at very low concentrations (∼200–400 NPs per L). Donovan et al. studied the presence of TiO₂, Ag and AuNPs,²⁴  and of ZnO and CeO₂ NPs,²⁵  in the source water (Missouri River) from three larges drinking water treatment facilities (DWTFs). Ti-Containing NPs, together with dissolved Ti, were detected in the river water samples, while Ag- and Au-bearing NPs were not detected.²⁴  On the other hand, Zn-containing and Ce-containing NPs were found in the source water from each facility.²⁵  Most metal-containing NPs were reduced by >95% from source water to finished drinking water. Ag-, Ti- and Ce-bearing NPs were detected in the Besòs River Basin.²⁶  A high concentration of Ti-bearing NPs (up to 298 × 10⁶ Ti-NPs per L) was detected in areas with little anthropogenic pressure. By determining the Ce/La ratio by single particle (SP)-ICP-MS, the authors concluded that the concentrations of Ce-bearing NPs found in river waters (from 18.1 × 10⁶ to 278 × 10⁶ NPs per L) were related to the natural occurrence of the mineral Monazite.

In a comparative study between measured and modelled concentrations, de Klein et al. investigated the occurrence of <450 nm Ce-, Al-, Ti- and Zr-based particles in River Dommel (The Netherlands).²⁷  The measurements, performed by asymmetric flow field-flow fractionation (AF4), confirmed the occurrence of the elements Ce and Al in the <450 nm sized particles in all water samples. Although Ti-bearing NPs were not directly detected, the authors suggested the presence of particulate TiO₂ based on its extremely poor solubility and the <450 nm total concentrations measured. Finally, Zr was assumed to be present as ZrO₂. In 2018, Markus et al. investigated the occurrence of metal-based NPs in the same river via ultrafiltration, high resolution (HR)-ICP-MS and scanning electron microscopy (SEM).²⁸  Authors claimed that 80–90% of the mass concentration of Ti in river samples was attached to natural colloids or was present as individual particles or clusters of smaller particles. In addition, the ratio of La and Ce concentrations indicated that all Ce was of a natural origin.

Gondikas et al. studied the occurrence of TiO₂ NPs in surface waters in a one-year survey at the Old Danube Lake (Vienna, Austria).²⁹  Ti-bearing NPs were identified during the whole 12 months period, with a concentration peak during the summer season, due to the release of this kind of NP from sunscreens. However, the techniques used in this first study were not able to discriminate sunscreen NPs from natural Ti-containing NPs. In 2018, thanks to the improvement and advances in analytical instrumentation, the same group used SP-ICP-MS and single particle ICP time-of-flight MS (SP-ICP-ToF-MS), which allows a multi-element quantification of individual particles, combined with microscopy techniques to distinguish between natural and engineered TiO₂ NPs in the same samples studied in the previous work.³⁰  While no specific multi-element signatures were detected for ENPs, the multi-element analysis of individual particles revealed that Al, Fe, Mn, and Pb were often present in natural Ti-bearing NPs.

On the other hand, urban stormwater ponds can also receive ENPs from multiple sources like vehicle emissions or painted façades.¹³  Baalousha et al. collected water samples from stormwater ponds within the city limits of Durham (USA) and studied the occurrence of different NPs. Nanosized FeO_x, TiO_x NPs, ZnO NPs, ZnS NMs and CePO₄ NMs were identified in the urban pond samples.

The emission of ENPs from wastewater treatment plants (WWTPs) may result in an increase of their concentration in certain river hotspots. For instance, Li et al. studied two representative full-scale municipal WWTPs next to the River Isar in Germany, showing the influence of the effluents on the AgNP concentration in the river.³¹  The authors found an increase in AgNPs levels up to 2.0–8.6 ng L⁻¹ due to the WWTPs, while concentration levels decreased at sites ∼1.5 km downstream of each discharge point. Similar results were found by Sanchis et al.²⁶  in a study performed in the Barcelona catchment area and in the Ebro River Delta. Ag-bearing NPs were only found at certain hotspots close to WWTP discharge points. On the other hand, Shi et al. found Ti-containing spherical nanoparticles with diameters of about 20–50 nm in the receiving waters (Xiaohe River) of two full-scale WWTPs.³² 

With respect to plastic particles, the presence of microplastics in marine and freshwater systems is well documented.^33,34  For instance, microplastics have been found in the Southern Ocean,³⁵  South Atlantic Ocean,³⁶  Latin American marine and freshwaters,³⁷  the South American Pantanal,³⁸  the coastal surface waters around the subtropical island of Okinawa (Japan),³⁹  subtropical gyres,⁴⁰  the Western Mediterranean Sea,⁴¹  the Sea of Marmara (Turkey),⁴²  the Yangtze Estuary System (China),⁴³  and in a continental environment such as the Greater Paris (France),⁴⁴  among others. However, the detection and quantification of nanoplastics is more challenging due to their small size. The occurrence of nanoplastics in real environment samples was demonstrated for the first time in 2017 by Ter Halle et al. in the North Atlantic subtropical gyre.⁴⁵  Authors examined four fractions (meso-, large micro-, small micro-, and nanoplastics) of the debris collected. Nanoplastics were obtained by isolating the colloidal fraction of seawater and further detected by dynamic light scattering (DLS). More recently, Devranche et al. identified the presence of nanoplastics on the beach exposed to the North Atlantic Gyre.⁴⁶  Authors determined the composition and size characterisation of nanoplastics by using transmission electron microscopy (TEM), DLS and pyrolysis gas chromatography MS (Py-GC–MS). In the next few years, the development of new strategies (including sampling protocols, separation methods and detection techniques) will allow the assessment of the presence of nanoplastics in water bodies.

WWTPs are a potential entry point of ENPs in the environment, despite the new findings and developments in the removal technology of ENPs in drinking water and wastewater treatment processes.⁴⁷ 

Roughly spherical TiO_x NPs with sizes ranging from 4 to 30 nm were found in wastewater effluents from 10 municipal facilities in the US.⁴⁸  Authors isolated the colloidal material via rota-evaporation, dialysis and lyophilisation. Detection and morphological characterisation was performed by high resolution (HR)-TEM and energy dispersive X-ray analysis (EDX). Colloidal Ag (defined as particles with sizes between 2 and 450 nm) and particulate Ag (≥450 nm) were found in nine separate sewage treatment plants in the UK.⁴⁹  The presence of Ag-bearing NPs, together with Ti- and Ce-bearing NPs, was also observed by SP-ICP-MS in the effluent of a WWTP in Gothenburg, Sweden.⁵⁰  Concentrations of 9568, 2312 and 32 656 particles per mL were obtained for Ag-, Ce- and Ti-bearing NPs, respectively. The presence of Ce-bearing NPs was explained by road runoff, and industrial and domestic wastewater ending up in the wastewater system, while Ti-bearing NPs were assumed to come from sunscreens and other cosmetics. However, the authors did not rule out natural sources for both metal-containing NPs. Ag-, Ce- and Ti-bearing NPs were also detected and quantified in influents and effluents of WWTPs in Catalonia, Spain.²⁶  The highest concentration in influent was found for Ti-bearing NPs (442 × 10⁶ NPs per L), followed by Ag-bearing NPs (149 × 10⁶ NPs per L) and Ce-bearing NPs (58 × 10⁶ NPs per L). The concentrations of these three metal-containing NPs showed elimination rates of 76, 76 and 66% for Ti, Ag and Ce NPs, respectively. High removal rates of ENPs in WWTPs have also been found in other studies. Li et al. demonstrated removal rates of >94.6% for AgNPs in representative WWTPs in Germany.³¹  Cervantes-Avilés and Keller studied the incidence of NPs from 13 different elements throughout the wastewater treatment process.⁵¹  Authors found removal rates ranged from 70 to 78% for Mg, Ni and Cd-based NPs, while 84–99% of natural and engineered metal-based NPs were removed from influent to reclaimed water. Similarly, Huang et al. showed that the removal efficiency of NPs was strongly dependent on element composition.⁵²  In this latter study, particle with sizes <100 nm from nine elements (Ag, Al, Au, Cd, Ce, Co, Cu, Ni and Zn) were detected in wastewater.

Vogt et al. assessed the presence of Ag-containing NPs in the influent of a WWTP discharging purified wastewater directly into Lake Mondsee (Upper Austria) over two years.⁵³  In this study, Ag-containing NPs were detected by SP-ICP-MS in the eight sampling campaigns. While the majority of the particles exhibited sizes close to the limit of detection of the technique (∼18 nm) some samples contained particles with sizes up to 98 nm. Although the authors did not report nanoparticle number concentrations due to the potential source of errors, the low particle per measurement time observed suggested that only a relatively small fraction of Ag-containing NPs enter the lake and the majority of the particles are effectively removed by the WWTP.

One of the main challenges when facing environmental pollution is the coexistence of the analyte of interest in its dissolved and in its nanoparticulate form. In this context, Mitrano et al. achieved the detection and quantification of the particulate and dissolved form of Ag in the same influent and effluent samples from a WWTP.⁵⁴  200 ng L⁻¹ Ag-bearing particles in the presence of 520 ng L⁻¹ dissolved Ag and 100 ng L⁻¹ Ag-bearing particles in the presence of 60 ng L⁻¹ dissolved Ag were quantified in the influent and effluent samples, respectively. In a similar way, Peters et al. detected the occurrence of 40 nm AgNPs at a concentration of 50 µg L⁻¹ in the presence of 450 µg L⁻¹ of dissolved Ag in a wastewater sample.⁵⁵ 

The daily use of consumer products based on NMs may become a potential source of ENPs entering the environment via wastewater. For instance, Farkas et al. analysed the effluent of a nanosilver producing washing machine.⁵⁶  Authors confirmed the presence of AgNPs at a concentration of 8 × 10⁷ mL⁻¹ by SP-ICP-MS and ion selective electrode measurements. The AgNPs detected were sized at an average diameter of 10 nm by TEM and 60–100 nm by nanoparticle tracking analysis (NTA).

Concerning carbon-based NMs, Farré et al. developed a method based on liquid chromatography (LC) coupled to a hybrid triple quadrupole linear ion trap mass spectrometry (QqLIT-MS) for the determination and quantification of fullerenes (C₆₀ and C₇₀) and functionalized fullerenes (N-methylfulleropyrrolidine C₆₀) on the suspended material of wastewater effluents.⁵⁷  The extraction from suspended solids in wastewater prior to the analysis was carried out by ultrasonication. This method allowed the detection and quantification of fullerenes in half of the effluents analysed of 22 WWTPs in Catalonia. On the other hand, Chen and Ding reported the presence of C₆₀ aggregates in industrial effluents and in surface waters in Taiwan, while C₇₀ was also found in industrial and municipal effluents.⁵⁸  Finally, Bäuerlein et al. investigated the presence of carbon-based NPs in different types of waters in The Netherlands.¹⁴  By using an HPLC system coupled to a hybrid LTQ Orbitrap mass spectrometer, the authors observed the presence of C₆₀ in all the six sewage treatment plants studied and in four of the effluents. However, C₇₀ was not detected in any of the influents or effluents. In addition, none of the fullerenes targeted was detected in dune and bank filtrates, surface waters and ground waters.

Finally, leachates from landfills for incineration residues should be considered as another potential pathway through which ENPs can enter the environment. The presence of colloids associated with different metals has already been observed in landfill leachates from municipal solid waste.^59,60  In this context, Mitrano et al. collected the leachate from a Swiss landfill for municipal solid waste incineration and studied the presence of ENPs.⁶¹  Anthropogenic TiO₂ NPs were identified in every leachate sample. In addition, the authors also observed discrete nanosized Ag/Au and Cu particles, although the anthropogenic origin of these NPs could not be confirmed since particle formation during the waste incineration process should not be discarded.

Some studies have reported the migration of NPs from water pipework material to tap water. For instance, lead dioxide (PbO₂) NPs can be found in lead-containing plumbing materials exposed to drinking water containing chlorine.⁶²  The formation of immobilized PbO₂ NPs has been suggested to be the result of transformations of soluble Pb(iii) species.⁶³  In a similar way, the use of Cu, which is the most common water pipe material in household drinking water pipeline systems, may lead to the release of Cu-bearing NPs into the water stream.⁶⁴  In addition, Cu always contain traces of Ag.⁶⁵  In this context, Wimmer et al. showed that Cu drinking water pipes were capable of releasing Ag ions as well as Ag-b-NPs into tap water that passes through copper pipes.⁶⁶  By using SP-ICP-MS, they concluded that 96% of the measurable particle sizes ranged from 10 to 36 nm. The percentage of Ag in nanoparticulate form with respect to the total Ag detected was reported at 30%. Authors showed with this study that tap water might be a source of Ag-bearing NPs in aquatic environments.

On the other hand, nanoplastics have been recently detected in tap water in Hangzhou (China).⁶⁷  Nanoplastics were successfully separated from tap water by sequential filtration and sizes determined by DLS. The most frequent particle sizes found were 255, 148 and 58 nm, at concentrations in the range 1.67–2.08 µg L⁻¹.

One of the environmental exposure routes of ENPs is the spreading of sewage sludge from WWTPs on soils, incineration plants or landfills.⁶⁸  Keller et al. estimated that, in 2010, 63–91% of the global ENM production ended up in landfills, followed by soils (8–28%).⁶⁹  For instance, it has been estimated that 36% of total emitted TiO₂ NPs ends up in soils via the deployment of sewage sludge, whereas 30% results from the deposition of sewage sludge into landfills directly or after incineration.⁷⁰  Therefore, the detection and quantification of ENPs in these environmental compartments is of the utmost priority.

Different ENPs have been found in sewage sludges. For instance, nanosized silver sulphide (α-Ag₂S) particles were identified by HR-TEM in the final stage sewage sludge of a full-scale municipal WWTP.⁷¹  The authors found loosely packed aggregates of ellipsoidal Ag₂S nanocrystals of 5–20 nm, which are likely formed in situ due to the reaction of reduced S with AgNPs or soluble Ag species, favoured by the anaerobic conditions during wastewater treatment. Similarly, Shi et al. found that 74–85% TiO₂ NPs in the influent sewage of two full-scale WWTPs in northern China was removed by the activated sludge.³²  However, authors observed significant Ti concentrations (27–43 µg L⁻¹) that remained associated with colloids in the effluents.

The mass concentration of nanoTi was determined at each stage process of a wastewater treatment process in a field-scale study at a municipal WWTP in central Arizona (US).⁷²  Although the authors found that the majority of Ti in raw sewage presented sizes >0.7 µm, and this fraction was well removed by the WWTP process, Ti concentrations ranged from <5 to 15 µg L⁻¹ in the <0.7 µm fraction were found in effluents. In addition, single NPs and aggregates composed of (presumably) TiO₂ < 50 nm spheres were observed in all samples. On the other hand, the distinction between natural and engineered TiO₂ NPs present in sanitary sewer overflows (SSOs) was studied by Loosli et al.⁷³  By using the Ti/Nb ratio, they demonstrated that SSOs are hot spots of engineered TiO₂ NPs released into the environments. Concentrations of up to 100 µg L⁻¹ were found in SSO impacted surface waters.

The transfer of TiO₂ NPs from sludge to soil was demonstrated by Kim et al.⁷⁴  By using SEM, TEM and scanning transmission EM (STEM) they identified TiO₂ NPs with sizes from 30 to 400 nm in three sewage sludge types with the same rutile structure as the TiO₂ NP aggregates found in surface soil mesocosms.

By using a sequential extraction method with acetic acid coupled with SP-ICP-MS, Tou et al. studied the presence of metal-containing NPs in sewage sludge samples collected from 26 WWTPs in Shanghai.⁷⁵  In combination with electron microscopy techniques, billions of Ti-, Fe-, Zn-, Sn- and Pb-containing NPs were detected in the acid-exchangeable fraction. The authors highlighted the fact that these NPs are likely to be released into the environment under acidic environmental conditions. The presence of ENPs has also been detected in sediments close to WWTP discharge points. For instance, AgNPs were detected in suspension after agitation of sediments from two sites in western Lake Ontario impacted by discharges from WWTPs serving the City of Toronto.⁷⁶ 

Soils amended with biosolids can accumulate metals in their ionic and nanoparticulate form after their concentration in WWTPs. Yang et al. studied the presence of metals and metal NPs in agricultural sites with long-term biosolid application.⁷⁷  Ti-Containing particles with sizes comprised between 50 and 250 nm were identified by TEM-EDX.

Plastic debris can be incorporated into soils through sewage sludge products. The identification of nanoplastics in soils is hampered not only by their size but also due to the amount of natural organic matter (NOM), which is also made of carbon. In 2021, Wahl et al. provided evidence of the occurrence of nanoplastics in agricultural soils collected in central France.⁷⁸  By using innovative analytical strategies, the authors confirmed the presence of nanoplastics composed of polyvinyl chloride (PVC), polystyrene (PS) and polyethylene (PE) with sizes ranging from 20 to 150 nm.

Once in the environment, ENPs can enter and accumulate in living organisms through different routes: through the lungs by inhalation; through the digestive tract by intake of water and food containing NPs; through the skin and mucous membranes; through exposure to contaminated surfaces; through the gills into the circulatory system of aquatic organisms.⁷⁹  Among all these exposure routes, the main ones are the gastrointestinal tract and respiratory tract.⁸⁰  In addition, ENPs can be transported through the whole food chain.

For instance, ENPs that end up in soils due to their use in different agricultural applications may be taken up by edible plants that are part of the human diet. The foliar application of nanofertilizers is another potential route of entry for these emerging contaminants. Although the occurrence of ENPs in plants has not been observed yet, the uptake of different kinds of metal-containing NPs with different sizes by edible plants and their transport to above ground organs has been demonstrated.⁸¹  Over the last few years, several studies have been performed in order to investigate the accumulation and effects of ENPs on aquatic and terrestrial organisms.⁷⁰  However, all these studies were carried out by preparing spiked samples or through in vivo exposure experiments.⁸² 

In recent years, micron and sub-micron plastics particles have been found in living organisms. For instance, the occurrence of microplastics has been reported in semiplagic fish (Boops boops) in the Balearic Islands (Mediterranean Sea) in a study carried out from March to May 2014.⁸³  The authors found microplastics in 68% of full stomach samples with a high frequency of occurrence (ranging from 42 to 80%) compared to other ingested items. In a study conducted in the seawaters of artificial reefs in Shengsi (East China Sea), Wu et al. reported the presence of microplastics in commercial fishes.⁸⁴  Microplastics were observed in 37.6% of fishes analysed, with fibre representing the 90.74% of plastics particles found. Additionally, the most abundant types of microplastics identified were cellulose and cellophane. Microalgae have also shown a strong ability to enrich microplastics from seawater. Gao et al. found microplastics in the size range 2–5 mm trapped by Ulva prolifera, a green tide species, from the Yellow Sea of China.⁸⁵  The main polymer types found were PS and polyethylene terephthalate (PET). Microplastics have also been found in zooplankton,⁸⁶  echinoderms,⁸⁷  cetaceans,⁸⁸  filter-feeding bivalves^89,90  and diving seabirds.⁹¹ 

Nanoplastics are also likely to be present in living organisms but their occurrence might be overlooked due to technical limitations. In 2018, Dawson et al. evidenced the fragmentation of microplastics (31.5 µm) ingested by Antarctic krill (Euphasia superba) into pieces less than 1 µm in diameter.⁹² 

This chapter has summarised the principal studies concerning the presence of ENPs in four of the main environmental compartments (air, water, soil and living organisms). Although remarkable improvement has been shown in the past decade, the study of the occurrence of ENPs in real environmental samples is still a challenge for environmental science for several reasons. On the one hand, the nanoparticle concentrations expected in the environment are below the limit of detection of the majority of techniques, even though the development of new analytical techniques and/or the technological development of many other ones, is helping to close the gap. On the other hand, ENPs are virtually present in all the environmental compartments which implies dedicated sampling procedures and sample treatments for each kind of matrix. On top of that, NPs are a special case of analytes, with very specific characteristics, which require a certain amount of information for their complete analysis and characterisation (e.g. size, shape, concentration, elemental composition, etc.). In addition, the presence of naturally occurring NPs in different environmental compartments may hamper the detection of anthropogenic ones, producing artefacts and hence unreliable results. Last but not least, an even bigger challenge is coming: the occurrence of nanoplastics in the environment, which cannot be analysed by simply cloning-off the current analytical techniques used for the characterisation of ENPs.