A review of microplastic removal from water and wastewater by membrane technologies Open Access

Microplastics (MPs) are plastics less than 5 mm in size and are classified as primary and secondary MPs based on their source (Acarer 2023b). MPs released from primary and secondary sources are present in surface waters (Egessa et al. 2020), groundwaters (Samandra et al. 2022), wastewaters (Franco et al. 2021), tap waters (Tong et al. 2020), and bottled waters (Mason et al. 2018) in different polymer types, shapes, sizes, and colors. Some of the MPs in the influent of conventional drinking water treatment plants (DWTPs) and wastewater treatment plants (WWTPs) are removed from the water/wastewater by passing through a series of treatment units in the DWTPs/WWTPs. However, MPs are still present in the effluents of both DWTPs and WWTPs (Ziajahromi et al. 2017; Dalmau-Soler et al. 2021; Franco et al. 2021; Sarkar et al. 2021; Acarer 2023b). Therefore, there is a need to incorporate existing treatment processes/technologies with a high removal efficiency of MPs into conventional DWTPs/WWTPs or to develop new technologies that separate MPs from water/wastewater with high efficiency.

Microfiltration (MF), ultrafiltration (UF), nanofiltration (NF), and reverse osmosis (RO) membranes, which are pressure-driven membranes, are advanced treatment technologies that are widely used in water and wastewater treatment. The overall MP removal efficiency of conventional DWTPs/WWTPs increases with the incorporation of membrane technologies into DWTPs/WWTPs (Dalmau-Soler et al. 2021; Cai et al. 2022). In addition, since the membranes have much smaller pore sizes than MP, they outperform the MP removal efficiency of different tertiary treatment technologies in WWTPs. Talvitie et al. (2017) reported that primary effluent-treating MBR reduced MP concentration from 6.9 ± 1.0 to 0.005 ± 0.004 MP/L. Moreover, it has been reported that the removal efficiency of MBR is higher than the MP removal efficiency of both conventional activated sludge (CAS) process and treatment units used as tertiary treatment (rapid sand filtration, dissolved air flotation, and disc filter) in different WWTPs and that the most effective technology in MP removal from wastewater is MBR technology (Talvitie et al. 2017). Cai et al. (2022) determined the removal efficiency of MPs in WWTP, which consists of two phases: phase 1, where primary sedimentation, biological treatment, and secondary sedimentation are applied, and phase 2, where primary sedimentation, biological treatment, MBR, UF, and RO are applied. Cai et al. (2022) showed that the overall MP removal rate in phase 2 where membrane technologies were applied was higher than in phase 1 where CAS was applied, and MP removal efficiency after MBR and RO were 93.2 and 98%, respectively (Cai et al. 2022). Similarly, Dalmau-Soler et al. (2021) compared the MP removal efficiency in DWTP, which is divided into two parallel lines after the treatment of lake water with coagulation, flocculation, sedimentation, and sand filtration. Dalmau-Soler et al. (2021) reported that the line with UF and RO had higher MP removal efficiency than the line with granular activated carbon (GAC) and ozonation.

This study aims to review the removal of MPs from water and wastewater by membrane technologies. Firstly, the abundance and properties of MPs in the influent of conventional DWTPs and WWTPs, as well as the removal efficiencies of DWTPs/WWTPs are reviewed. Then, studies examining MP removal by MF, UF, NF, RO, and MBRs and the effect of MPs on membrane fouling are summarized. After that, the advantages of MP removal from water/wastewater by membrane technologies compared to other removal methods are discussed. Moreover, studies on MP release into water/wastewater from polymeric membranes are discussed. Finally, studies on the fouling of membranes by MPs are reviewed, and recommendations are made for reducing membrane fouling in future studies. It is worth mentioning that, to the best of my knowledge, there is no study in the literature that reviews the properties of MPs to which membranes used in water/wastewater treatment are exposed, MP removal efficiencies of membranes, fouling of membranes with MPs, and MP release potential from membrane structures to water/wastewater.

Since membrane technologies are advanced treatment technologies that provide high-efficiency contaminant removal, they are used in WWTPs and DWTPs to provide better quality treated water in water and wastewater treatment. Reviewing the abundance and characteristics of MPs in different WWTPs and DWTPs is useful for better understanding the amount and properties of MPs that membrane technologies frequently encounter in water and wastewater treatment, improving treatment technologies, and taking necessary measures. Table 1 summarizes the results of some studies on the abundance, polymer types, shapes, and sizes of MPs in influents of WWTPs/DWTPs.

The abundance of MPs in the influent of WWTPs and DWTPs varies depending on the density of the population in the area where the water/wastewater is collected and the habits of the population, waste management strategies, urbanization, traffic density, and seasonal conditions (Gao et al. 2022; Uogintė et al. 2022). Currently, there is no standard analysis method for determining the amount of MP in drinking water and wastewater. The fact that the water/wastewater sample volumes collected, the methods and devices used, and the size range of the MPs examined in the studies are quite different makes it complex to directly compare the studies with each other.

In the influents of WWTPs and DWTPs, especially fiber and fragment-shaped MPs are dominant (Table 1). Fibers are released into wastewater mainly as a result of washing synthetic clothes in the washing machine. It has been reported that 137,951; 496,030; and 728,789 fibers are released in one laundering of 6 kg polyester (PEST)-cotton blend, PEST, and acrylic synthetic clothing, respectively (Napper & Thompson 2016). As a result of the discharge of treated wastewater into surface waters, the fibers enter drinking water sources and contribute to the abundance of the influents of DWTPs. It has been reported that the dominant MP shape in effluents of WWTPs treated by passing through pre-treatment, primary treatment, secondary treatment, and tertiary treatment units is fiber (Mason et al. 2016; Franco et al. 2021). As a result of the discharge of treated wastewater into surface waters, fibers enter drinking water sources and contribute to fiber abundance in the influent of DWTPs. Fragmented MPs in water sources are an indication that macro and mesoplastic wastes are fragmented into smaller pieces to form MPs.

Generally, very small-size MPs (<10 μm) dominate in the influent of DWTPs (Pivokonsky et al. 2018; Wang et al. 2020; Adib et al. 2021; Shen et al. 2021; Sharifi & Movahedian Attar 2022) while MPs >100 μm in size dominate in the influent of WWTPs (Gündoğdu et al. 2018; Long et al. 2019; Franco et al. 2021). Conventional DWTPs and WWTPs still have MPs in their effluents, even if the water/wastewater passes through a series of treatment units. MPs in the effluent of DWTPs reach consumers through tap water and pose a possible health risk to humans. Considering the amount of MP in 1-L effluent of WWTPs (MP/L) and the WWTP capacity, millions of MPs are released to the receiving environments in 1 day (Acarer 2023b). MPs released to the receiving environment adsorb organic and inorganic contaminants (Guo et al. 2019; Sharma et al. 2020; Ta & Babel 2020; Mo et al. 2021), accumulate in the bodies of organisms in the aquatic environment, and pose a health risk by being transferred to other organisms through the food chain (Alfaro Núñez et al. 2021).

As a result, more removal of MPs in different polymer types, shapes, and sizes in conventional DWTPs and WWTPs for human and environmental health, needs to be developed, and/or existing technologies that exhibit high removal efficiency should be included in conventional DWTPs/WWTPs. Membrane technologies are a good option to make water and wastewater more suitable for human and environmental health by removing more MPs from water and wastewater that conventional treatment processes cannot remove.

Since inorganic membranes have better thermal, chemical, and mechanical resistance compared to polymeric membranes, they can be used even under harsh environmental conditions, unlike polymeric membranes. In addition, inorganic membranes have high flux due to their high hydrophilicity (low contact angle) and better separation efficiency and antibacterial property due to narrow pore size distribution (Hofs et al. 2011; He et al. 2019). Moreover, since inorganic membranes have a longer lifespan (>10 years) than polymeric membranes (up to 10 years), they serve longer treatment periods. Although inorganic membranes have many advantages over polymeric membranes, inorganic membranes have two major disadvantages due to their brittleness and high production cost. Since these two major disadvantages make the production and assembly of inorganic membranes more difficult, polymeric membranes are still more widely used today both in large-scale applications such as treatment in DWTPs/WWTPs and in small-scale applications such as laboratory studies.

The membranes used in water/wastewater treatment should have a low cost, high mechanical, chemical, and thermal resistance, and a low fouling tendency. Therefore, to obtain superior properties and performance in polymeric membranes, researchers continue to study determining the appropriate amount of polymer (Alvi et al. 2019), amount of solvent and type (Acarer 2022; Yin et al. 2022), and amount of pore-forming agents (Gebru & Das 2017) during membrane production. In addition, in the current situation, researchers focused on improving the properties of the membrane and increasing the flux and separation performance by adding very low amounts of different inorganic nanomaterials into the matrix of polymeric membranes (Polisetti & Ray 2021; Srivastava & Raval 2022). In most of the studies in which polymer-based nanocomposite membranes are produced, researchers have investigated the effects of nanomaterial type, nanomaterial amount, nanomaterial modification, and production conditions on the membrane properties and performance (Nikita et al. 2019; Poon et al. 2023). Increasing studies on the production and comprehensive characterization of polymer-based membranes can contribute to the determination of the appropriate membrane composition to bring the very advantageous aspects of ceramic membranes to low-cost polymeric membranes.

In this section, studies on MP removal from waters and wastewaters by MF, UF, NF, RO, and MBRs are reviewed. The results of some studies on the MP removal efficiency of membranes used in water and wastewater treatment are summarized in Table 2.

MF membranes are pressure-driven porous membranes with a pore size of 0.1–10 μm and operate in the range of 0–2 bar (Moradihamedani 2022). MF is generally used for the removal of suspended solids, colloids, and bacteria from water and wastewater. In addition, in the treatment of water and wastewater, MF membranes are preferred as a pre-stage to reduce clogging and/or fouling before pressure-driven membranes with smaller pore sizes (UF, NF, and RO).

While it has been revealed in the studies that the MP removal efficiency of MF membranes in DWTPs/WWTPs varies between 81.5 and 100%, laboratory studies have shown that the MP removal efficiency of MF membranes up to 98.5% (Table 2). MPs with a larger size than the pore size of the MF membrane are separated from the water/wastewater by being retained on the membrane surface and/or pores. The size of MPs in treated waters of conventional DWTPs is generally <100 μm and studies show that conventional DWTPs are not very effective in removing small-sized MPs (Pivokonsky et al. 2018; Adib et al. 2021; Kankanige & Babel 2021; Shen et al. 2021; Sharifi & Movahedian Attar 2022). Similarly, the dominance of MPs smaller than 500 μm in the effluent of WWTPs indicates that small MPs are not effectively removed in WWTPs, and especially small MPs are discharged into the receiving aquatic environments (Long et al. 2019; Tadsuwan & Babel 2021). The inclusion of membranes with MF characteristics in conventional DWTPs and WWTPs also contributes to the higher efficiency removal of small MPs from water and wastewater that cannot be effectively removed in conventional DWTPs/WWTPs. On the other hand, in the study by Pizzichetti et al. (2021) it was reported that although the pore size of MF membranes was smaller than the size of MPs, MPs were found in the permeate of the MF membrane after filtration and 100% removal efficiency could not be achieved. In the study, it was suggested that low membrane strength and sharp-cornered MPs may cause abrasion of the membranes among the reasons for MPs to pass through the membranes and reach the permeate (Pizzichetti et al. 2021). In order to achieve effective MP removal by MF membranes, the membrane pore size must be smaller than the lower limit value of the MP size range, i.e., 1 μm, and the mechanical strength of the membrane must be resistant to operating pressure, and the physical properties of the membrane. It is also an important factor that the tank in which the permeate is collected is a closed tank, and the tank is designed in a way that does not contribute to MP release. It should also be ensured that no MP is released into the water/wastewater from the structure of the polymeric membranes.

UF membranes are typically porous membranes with a pore size in the range of 1–100 nm and operated at a pressure of 1–10 bar (Karimi et al. 2019). UF membranes are generally used for the removal of suspended solids, colloids, bacteria, viruses, and high molecular weight organic compounds from water/wastewater (Jacquet et al. 2021; Tomaszewska & Mozia 2002). Since the average pore size of UF membranes is smaller than 100 nm, they provide superior performance in the removal of MPs (1–5,000 μm) from water/wastewater. For instance, Ma et al. (2019) reported the complete removal of PE MPs of different sizes (<500–5,000 μm) in deionized water by a commercial PVDF flat sheet membrane with an average pore size of 30 nm. In a study by Tadsuwan & Babel (2022), as a result of treating wastewater in conventional municipal WWTP (coarse screen, fine screen, grit chamber, aeration tank, and final clarifier), MPs in the influent (77 ± 7.21 MP/L) were removed with 86.14% efficiency and decreased to 10.67 ± 3.51 MP/L in the effluent. By feeding the effluent of the final clarifier to the UF membrane, the UF membrane reduced the MP concentration to 2.33 ± 1.53 MP/L with a removal efficiency of 78.16%, and the UF membrane increased the MP overall removal efficiency of conventional WWTP to 96.97% (Tadsuwan & Babel 2022).

Since the pore size of UF membranes is smaller than that of MF membranes, MP removal by UF membranes is higher than that of MF membranes. Luogo et al. (2022) investigated the removal of contaminants from laundering wastewater using a UF membrane made of ZrO₂ and an MF membrane made of SiC. The results of the study revealed that the UF membrane performed better in removing total suspended solids, volatile suspended solids, chemical oxygen demand, turbidity, and MP from wastewater due to the smaller pores in the selective layer of the UF membrane compared to the MF membrane. In the study, the MP removal efficiency was found to be 98.5 and 99.2% by MF and UF membranes with 90% of the pores smaller than 302 and 63 nm, respectively (Luogo et al. 2022). Similarly, Pramanik et al. (2021) also reported that the PES UF membrane removes nanoplastics and MPs with higher efficiency (up to 96%) compared to the PVDF MF membrane (up to 91%) since the pore size of the PES UF membrane is smaller than the pore size of the PVDF MF membrane.

On the other hand, although the reduction of the pore size in membranes is an important factor in increasing the MP removal efficiency, the material of the membranes must be the same to compare the MP removal efficiency of MF and UF membranes depending on the pore size. MPs have a high contact angle, i.e., MPs are hydrophobic (Ayush et al. 2022). Therefore, there are different interactions between hydrophobic or hydrophilic membrane surfaces and MPs. Since different membranes have different surface charges, repulsion due to the same charge or attraction forces due to opposite charges between the surface and MPs also affect the removal efficiency (Pizzichetti et al. 2021). Therefore, in future studies, there is still a need to comprehensively investigate and compare the removal efficiencies of MPs in different polymer types depending on the pore size, porosity, contact angle, zeta potential, roughness, and other properties of MF and UF membranes made of the same material. Thus, the interactions between membrane properties and MPs can be better understood, and improvements can be made to increase MP removal efficiency and/or reduce membrane fouling.

NF is a pressure-driven membrane with a pore size of 1–10 nm and operated at 5–15 bar. NF membranes are especially used for the removal of multivalent salts and organic molecules with molecular weight cut-off (MWCO) > 200 Da from water/wastewater (Puthai et al. 2017; Kang et al. 2020). NF membranes have the advantage of more salt rejection than UF membranes under the same operating conditions and have higher flux performance than RO membranes (Abid et al. 2012). However, due to the very small pore sizes of NF membranes, the high-pressure requirement for the passage of water through the membranes limits their widespread use. It has been reported that the use of NF membranes in water treatment increases the energy requirement for treatment by 60–150% (Abdel-Fatah 2018).

Studies examining the removal of MPs by NF membranes are very limited. Kara et al. (2023) reported that 96 and 99% of MPs were removed, respectively, after UF and NF processes in the landfill leachate treatment plant (LLTP) in Turkey. In the study by Kara et al. (2023), the abundance of MP in the leachate after the NF membrane was determined as 2 MP/L, and it was reported that the majority of MPs in the leachate after the NF process were larger than 500 μm and fiber shaped. Barbier et al. (2022) examined the abundance of MPs in a total of six samples collected before and after degassing following an NF membrane with an MWCO of 400 Da (pore size: 1 nm) in a DWTP located in France. In the study, MP could not be detected in four of the six samples, and 0.018 MP/L (before degassing) and 0.002 MP/L (after degassing) in the other two samples (Barbier et al. 2022). However, further investigation of MP removal from water/wastewater by nanofiltration membranes is needed in future studies.

RO membranes are dense membranes (pore size: <1 nm) (Acarer 2023b), these membranes are operated at high pressures (>20 bar) to ensure the passage of water through the membranes, and, as a disadvantage, their energy requirements are quite high. The separation process in RO membranes is based on the sorption-diffusion mechanism (Licona et al. 2018). RO membranes are used in the desalination of seawater and brackish water, as well as drinking water and wastewater treatment, as RO membranes can remove even monovalent ions in addition to all contaminants that MF, UF, and NF membranes can reject from water (Ansari et al. 2021; Indika et al. 2021).

RO membranes can be used in situations where contaminant removal is desired with higher performance than MF, UF, and NF membranes and where very strict limit values are applied in the effluent. It has been reported in studies that MPs are present even in the permeate of RO membranes used as an advanced treatment in WWTPs and DWTPs. For instance, Dalmau-Soler et al. (2021) revealed that membrane technologies (UF and RO) exhibit better MP removal than upgraded conventional treatment (ozonation and GAC filtration). However, even after RO, MPs were present (0.06 ± 0.04 MP/L) and about 60% of these MPs were fibers (mostly 500–2,000 μm) and about 40% were fragments (mostly 20–500 μm) (Dalmau-Soler et al. 2021). Similarly, Sun et al. (2021) reported that MPs <50 μm in size were encountered after the RO unit, and the majority of them were composed of fibers and the rest of fragments. Fiber-shaped MPs detected in the permeate of the membranes are an indication that the fibers can pass through the membranes longitudinally due to the high length-to-width ratio (Ziajahromi et al. 2017). Although it is seen that there is a low amount of MPs in the permeate of the membranes used in the treatment of wastewater, especially when the treatment capacities of WWTPs are taken into account, millions of MPs enter the receiving environment in 1 day. For example, Ziajahromi et al. (2017) reported that the MP abundance in the effluent of the RO system, which has a treatment capacity of 48 million liters per day, is 0.21 MP/L and approximately 10 million MPs are released from the RO effluent to the aquatic environment in 1 day.

However, it is interesting that fibers are present in the permeate even after water/wastewater treatment with the RO membrane, which is classified as non-porous and can even retain ions. It should be investigated in detail to determine whether the MPs present in the permeate of RO membranes are caused by passage through possible large openings in the membrane, release from the membrane structure to the water/wastewater, or environmental conditions.

MBRs have many advantages that can overcome the disadvantages of CAS systems, including eliminating the need for a secondary settling tank, providing higher-quality effluent, and producing less sludge (Yang et al. 2020). Moreover, MBRs allow operation at short hydraulic retention time, high mixed liquor suspended solid concentrations, high loading rates, and high sludge retention time (Barreto et al. 2017; Prasertkulsak et al. 2019). On the other hand, the high initial investment and operating costs of MBRs, membrane fouling, and replacement of the membrane after a while are the factors limiting the widespread use of MBRs (Rahman et al. 2023).

MBRs exhibit higher MP removal efficiency compared to CAS (Talvitie et al. 2017; Lares et al. 2018). In the secondary settling tank after biological treatment in CAS, the inability of MPs to settle well enough depending on their properties and WWTP conditions may cause MPs not to be removed very effectively (Kankanige & Babel 2021). MPs, which cannot be effectively removed due to poor settling in the CAS process, can be removed from wastewater with high-efficiency thanks to the membranes in MBRs (Talvitie et al. 2017). Talvitie et al. (2017) reported that MP concentration decreased from 6.9 to 0.005 MP/L (99.9% removal efficiency) as a result of the treatment of primary wastewater with the MBR pilot system using flat sheet membranes with a nominal pore size of 0.4 μm in a WWTP in Finland. In the study, it was determined that the MP concentration in the secondary wastewater in WWTP without pilot-scale MBR decreased to 0.2 ± 0.06 MP/L. In other words, the study found that the MP concentration in the permeate of MBR corresponds to a very small value (2.5%) of the MP concentration in the effluent of CAS (Talvitie et al. 2017). In the same WWTP in Finland, Lares et al. (2018) found MP concentration in wastewater passing through grit separation, primary clarification, and pilot MBR as 0.4 ± 0.1 and MP removal efficiency as 99.4%. Lares et al. (2018) also found MP concentration in wastewater treated with a CAS system as 1.0 ± 0.4 MP/L and MP removal efficiency as 98.3%. More MP accumulation in MBR sludge compared to primary/secondary sludge of WWTPs operated with CAS is also an indication that more MP is removed by MBR (Di Bella et al. 2022).

Studies have shown that MP properties (polymer type, shape, size, etc.), coagulant/flocculant type and dosage, operating conditions (mixing speed, pH, etc.), and the presence of other contaminants in the water are effective in MP removal by coagulation/flocculation (Lapointe et al. 2020; Na et al. 2021). Some studies have revealed that small-size MPs are removed with higher efficiency than large-size MPs by coagulation/flocculation/sedimentation process (Ma et al. 2019; Kankanige & Babel 2021). For instance, Ma et al. (2019) investigated the removal efficiency of PE particles classified as <0.5 mm, 0.5–1 mm, 1–2 mm, and 2–5 mm by coagulation/flocculation/sedimentation processes using FeCl₃·6H₂O and AlCl₃·6H₂O. The removal efficiency of <0.5 mm PE particles with 0.5 and 5 mM FeCl₃·6H₂O was found to be 8.24 ± 1.22% and 12.65 ± 1.09%, respectively, while the removal efficiency of <0.5 mm PE particles for 0.5 and 15 mM AlCl₃ 6H₂O 8.28 ± 1.06% and 36.89 ± 3.24%, respectively. In addition, it was determined that the removal efficiency decreased as the size of the PE particles increased (Ma et al. 2019). In contrast to the low MP removal efficiency with coagulation/flocculation, high MP removal efficiency is achieved for all MP types with membrane processes (Ma et al. 2019; Luogo et al. 2022; Kara et al. 2023). In addition, the fact that no chemicals are needed during the separation of MPs from water/wastewater by membrane processes, and the fast and simple operation of membrane processes make them more advantageous than coagulation/flocculation.

While the sedimentation process is generally applied after the coagulation/flocculation tanks in DWTPs, it is applied as primary sedimentation following the pre-treatment and secondary sedimentation following the biological treatment in WWTPs. MPs in water/wastewater settle together with other organic and inorganic contaminants in sedimentation tanks under the effect of gravity. As a result of MP removal studies conducted by Long et al. (2019) in different WWTPs in China, the mean removal rate for PET, PS, PE, and PP was determined to be 96.4, 94.8, 87.8, and 92%, respectively. In the study, PET MPs with higher density (1.37 g/cm³) than wastewater were easily removed by physical sedimentation. However, low-density PE and PP MPs (0.91 and 0.91–0.97 g/cm³) and moderate-density PS MPs (1.05 g/cm³) were removed with lower efficiency as they could float on the wastewater surface or remain suspended in the wastewater column (Long et al. 2019). Liu et al. (2019b) it was determined that MPs in the 0.02–0.3 mm size range were dominant in the effluent of the primary sedimentation, even if the wastewater was treated in pre-treatment and primary sedimentation units. Therefore, small MP size and low MP density have a negative effect on the removal efficiency of MPs by physical sedimentation (Long et al. 2019; Pittura et al. 2021). Moreover, MPs may not be effectively removed by sedimentation if sufficient time is not provided in the sedimentation tanks, as the settling of small-sized particles is slower. In the flotation process, which is the opposite of the sedimentation process, factors such as density, roughness, hydrophilic/hydrophobic properties, and interactions of MPs were found to be effective in removal (Wang et al. 2021). Whereas, since the pore sizes of membranes are significantly smaller than the size of MPs, all MPs can be effectively removed from water/wastewater, regardless of size, with membrane technologies. Moreover, all problems related to insufficient settling/float ability of MPs in settling or flotation units are eliminated by using membrane processes.

MPs can be effectively removed from water/wastewater by retention inside the sand filter media and deposition on the surface of the filter media (Talvitie et al. 2017). Depending on the characteristics of the sand filter media, large-size MPs are removed from water/wastewater with high removal efficiency, while small-size MPs are removed with lower efficiency (Wang et al. 2020; Babel & Dork 2021; Kankanige & Babel 2021). In addition, some studies have revealed that fiber-shaped MPs are removed with higher efficiency compared to other MP shapes by sand filtration (Wang et al. 2020; Babel & Dork 2021). The pore sizes of the membranes are much smaller than the pore sizes in the sand filtration medium. Thus, small MPs in water/wastewater that cannot be effectively removed by sand filtration can be effectively removed by membrane filtration.

Photolytic, chemical, photochemical, and photocatalytic advanced oxidation processes (AOPs) are used in the degradation of MPs. In a study by Easton et al. (2023), it was determined that the mass of PEST fibers in water decreased by 2, 6.3, and 10%, respectively, after 9 h of treatment with chemical oxidation with H₂O₂, photooxidation with UVC, and photochemical oxidation with UVC/H₂O₂. The study results showed that an AOP based on UVC/H₂O₂ was more successful in reducing the mass of PEST fibers faster than either UVC or H₂O₂ alone (Easton et al. 2023). In the study, it was determined by SEM analysis that the oxidation of 9 h treated PEST fibers with UVC/H₂O₂ made the surfaces of PEST fibers more rough compared to other oxidation methods and caused more pitting and hole formation on the surface of PEST fibers (Easton et al. 2023). Liu et al. (2019a) reported that the surfaces of smooth pristine PS and PE MPs became rougher after 30 days of heat-activated K₂S₂O₈ and Fenton treatment, and cracks and pits were created on the surfaces of MPs. In the study, due to the relatively stronger oxidation capacity of the heat-activated K₂S₂O₈ system compared to the Fenton system, the heat-activated K₂S₂O₈ system caused more observable changes on the surfaces of the MPs (Liu et al. 2019a).

Semiconductor metal oxides such as TiO₂ (Llorente-García et al. 2020; Vital-Grappin et al. 2021) and ZnO (Tofa et al. 2019; Uheida et al. 2021) have been used for the degradation of MPs by photocatalysis, which stands out as an environmentally friendly method. In the study of Llorente-García et al. (2020), the mass loss by photodegradation of 814 and 382 μm spherical HDPE MPs in 50 h visible irradiation using mesoporous N-TiO₂ coating was found to be 0.22 and 4.65%, respectively. Under the same conditions, the mass loss for 5 mm × 5 mm and 3 mm × 3 mm film LDPE MPs was 0.97% and 1.38%, respectively (Llorente-García et al. 2020). Uheida et al. (2021) reported that PP MPs decreased in size from 154.8 to 108.2 μm and the volume of PP MPs decreased by over 65% after exposure to 456 h of visible light irradiation using ZnO nanorods. In addition to the properties of the synthesized catalysis and MPs for the photocatalytic degradation of MPs, aquatic conditions such as temperature and pH also affect the degradation efficiency and speed of MPs (Ariza-Tarazona et al. 2020). Disadvantages of MPs removal from water by the photocatalysis process include prolonged exposure to visible light for degradation of MPs, low removal efficiency, and the need for catalysis synthesis. In addition, studies investigating the photocatalytic degradation of MPs in the literature mostly deal with PE MPs (Ariza-Tarazona et al. 2020; Jiang et al. 2020; Llorente-García et al. 2020). However, the effect of different photocatalysts on MPs with different properties should be investigated further, since water and wastewater media contain MPs in many different polymeric structures.

The general disadvantages of AOPs processes used in the degradation of MPs include the use of chemicals, energy, and catalysis. In addition, the degradation efficiency of MPs from water/wastewater by AOPs depends on properties such as polymer type, mechanical properties, and thickness of MPs (Liu et al. 2019a). As it is clear from the research studies mentioned above, the need for long times for the degradation of MPs by AOPs and the generally low removal efficiency are among the disadvantages. However, membrane processes are suitable for the rapid and high-efficiency removal of different MPs from water/wastewater, regardless of the properties of MPs.

Although polymeric membranes are widely used in water and wastewater treatment due to their low cost, easy production, good permeability, and good separation efficiency, several studies in recent years have suggested that polymeric membranes release MP into water/wastewater from their structure (Gan et al. 2021; Sun et al. 2021; Zhang et al. 2021). Zhang et al. (2021) found that while the MP abundance in leachate decreased gradually up to UF unit in LLTP, where adjustment tank, MBR, AO, UF, NF, and RO processes were applied, respectively, MP abundance increased after NF and RO. Although cellulose nitrate, which is the material of NF and RO membranes, was not encountered in the leachate after other treatment units in LLTP, it was found after NF and RO, which proved that MP has released from the structures of the membranes to leachate (Zhang et al. 2021). Gan et al. (2021) reported that as a result of ultrapure water filtration from commercial PVDF UF hollow fiber membrane (MWCO: 50 kDa, pore size: 0.02 μm), hydraulic blow on the membrane causes MP release. They detected 55.1 MP/L in the membrane effluent and reported that the majority of MPs detected in the effluent consisted of fragments (64%) and 1–10 μm (60%) MPs. Soaking and keeping the membranes in cleaning agents (citric acid, NaOH, NaClO) caused a change in the physical and chemical properties of the membranes and increased the number of MPs in the membrane effluent (Gan et al. 2021).

Membranes with different polymeric structures have different mechanical strengths and chemical resistances even if they have the same pore size and/or MWCO. Therefore, it is necessary to examine the release of MP into water/wastewater by filtration under different conditions from different polymeric membrane materials commonly used in water/wastewater treatment in future studies.

Laboratory-scale production, characterization, and application of polymeric membranes in water/wastewater treatment is a very popular study topic in recent years, which has become the focus of attention of researchers (Nasrollahi et al. 2018; Polisetti & Ray 2021; Acarer 2022). However, to the best of my knowledge, there is no study in the literature examining the release of MP into water/wastewater from membranes produced at a laboratory scale. In future studies, besides detailed characterization of the produced membranes, investigating possible MP release from membranes to water/wastewater and evaluating the results by establishing a relationship between membrane characteristics and release will help to better understand MP release conditions and to take necessary measures in DWTPs/WWTPs. To understand whether MP is released from the membrane to the water, the following procedure can be followed: (1) detailed characterization of the chemical structure of all polymeric materials of the membrane and the chemical structure of MPs in the inlet and outlet stream of the membrane by techniques such as FTIR, (2) comparison of the polymer types of the membrane and the polymer types of MPs in the permeate, and (3) investigation of the abundance of MPs of the same polymer type as the membrane polymer type in the permeate after membrane filtration (Acarer 2023a).

Transmembrane pressure (TMP) increase, and flux reduction are good indicators for understanding membrane fouling. TMP increases and flux decreases as contaminants accumulate in the membrane pores/surface. In addition to organic and inorganic contaminants in water/wastewater, MPs cause a further increase in membrane fouling. Therefore, the presence of MPs in water/wastewater causes a further decrease in the flux in the membranes and a greater increase in TMP. In the study of Enfrin et al. (2020), the flux decreased by less than 15% after 48 h of distilled water filtration at 1 bar through the UF PSf membrane, while the flux decreased by 38% with the addition of MPs/NPs to the feed under the same experimental conditions. The study of Li et al. (2021) showed that the presence of MPs in raw water, increasing MP concentration, and decreasing size of MPs (from 18 to 1 μm) increased membrane fouling. As a result of filtering the raw water through the PVDF hollow fiber UF membrane without and with PS MPs (1 mg/L, 1 μm), TMP increased to 35.5 and 74.0 kPa, respectively, on the 10th day (Li et al. 2021). Similarly, Larue et al. (2022) determined the average permeate mass of the membranes as 622 ± 55 g and 943 ± 70 g, after filtering the wastewater collected from the surface of the secondary clarifier tank of a WWTP in Canada through PVDF flat sheet UF membrane and MF membrane, respectively. Moreover, with the inclusion of 0.1, 1, and 10 mg/L PE MPs in the wastewater, the mean permeate mass of the UF membrane decreased to 483 ± 25 g, 352 ± 30 g, and 293 ± 15 g, respectively, while the average permeates mass of the MF membrane decreased to 709 ± 34 g, 644 ± 67 g and 322 ± 12 g. The fact that the decrease in the mean permeate mass of the MF membrane at the highest concentration of MP (10 mg/L) was higher than that of the UF membrane was associated with MPs clogging the MF membrane pores more easily due to the larger pore size of the MF membrane (Larue et al. 2022). In addition, studies have shown that small-size MPs cause more fouling of the membranes, contributing to a further decrease in flux and a greater increase in TMP (Ma et al. 2019; Li et al. 2021). There is still a need for further research with different membrane and MP types to determine the membrane materials, membrane properties, and operating conditions that allow the membranes to be used for MP removal for a longer time without fouling and compromising flux performance.

In some studies conducted in recent years, it has been reported that the incorporation of hydrophilic nanomaterials into polymeric membranes improves the membrane surface hydrophilicity as well as the antifouling property of the membrane (Wang et al. 2018; Chu et al. 2020; Behdarvand et al. 2021). The antifouling ability of the membranes is expressed by the flux recovery ratio (FRR), and the high FRR of the membranes indicates a high fouling resistance of the membrane. It has been reported that the FRR for contaminants such as oil (Ajibade et al. 2021), bovine serum albumin (BSA) (Ghezelgheshlaghi et al. 2018; Saraswathi et al. 2020; Ajibade et al. 2021), humic acid (Ghezelgheshlaghi et al. 2018; Park et al. 2020; Saraswathi et al. 2020), sodium alginate (Ghezelgheshlaghi et al. 2018), and dye (Koutahzadeh et al. 2016) increased after the incorporation of hydrophilic nanomaterials into pristine membranes. The increased membrane surface hydrophilicity weakens the interaction between the membrane surface and the contaminant and allows easy removal of deposited contaminants on the membrane surface by cleaning. However, there is no study in the literature investigating the FRR of the membrane after filtering MP-containing water/wastewater through membranes and then hydraulic/chemical cleaning of fouled membranes. In future studies, the FRR of MP-containing water/wastewater filtered membranes should be investigated by making connections between the surface properties of the membranes (hydrophilicity, roughness, charge, etc.) and the properties of MPs (polymer type, size, shape, hydrophobicity, roughness, charge, etc.). Thus, progress can be made in the fabrication of membranes with physical and chemical properties that are more resistant to MP contamination. Moreover, membranes resistant to MP contamination contribute to the reduction of the cleaning period and the amount of chemicals used in cleaning, thus reducing operating costs.

In some studies conducted in recent years, the self-cleaning ability of membranes for different contaminants has been investigated (Geng et al. 2019; Boopathy et al. 2020). Membranes with self-cleaning ability via photocatalytic degradation are promising filtration systems that use less chemical cleaning agent, longer process time, less frequent cleaning, and less process footprint (Lv et al. 2017; Bartot Coelho et al. 2021). When photocatalytic materials such as TiO₂ and ZnO included in the membrane are irradiated with light with higher energy than the band gap energy, electron and hole pairs are produced. While electrons react with oxygen molecules to produce superoxide radicals, holes react with water to produce hydroxyl radicals (Damodar et al. 2009; Al Mayyahi and Deng 2018). The generated radicals reduce the membrane contamination by decomposing organic and bioorganic foulants on the membrane surface and pores into H₂O and CO₂ or simple intermediates (Zangeneh et al. 2019). It has been reported that organic contaminants on the surface of catalytic membranes are more easily degraded as a result of exposure to UV light and their self-cleaning ability is higher compared to non-catalytic membranes (Zangeneh et al. 2019; Boopathy et al. 2020; Zhou et al. 2021).

On the other hand, although there are studies examining the degradation of MPs by photocatalysis, which is one of the AOPs (Jiang et al. 2020; Llorente-García et al. 2020; Uheida et al. 2021) no study examining the self-cleaning performance of the membrane for MPs by photocatalytic mechanism has been found in the literature. Application of membranes with self-cleaning ability in the filtration of MP-containing water/wastewater can provide a more efficient and longer-term treatment performance and reduce maintenance costs. For this reason, the removal of MPs from water/wastewater with self-cleaning membranes is worth considering in future research.

Literature analysis has shown that conventional DWTPs and WWTPs cannot completely remove MPs in drinking water and wastewater, and effluents of DWTPs/WWTPs contain MPs. On the other hand, the absence of a standard method for the analysis of MPs in water and wastewater makes it difficult to compare the results of studies in the literature with each other.

It can be said that membrane technology, which is one of the advanced treatment technologies for more effective removal of MPs from drinking water and wastewater, is the most effective and promising treatment technology. The fact that the pore sizes of the membranes are quite small compared to MPs and that the main mechanism in membranes is based on size exclusion makes them the ideal for MP removal. Therefore, the application of MF, UF, NF, and RO pressure-driven membrane technologies in DWTPs and WWTPs and MBRs in WWTPs reduces the MP concentration in the effluents of conventional DWTPs/WWTPs. Thus, membrane technologies allow less MP to reach tap water from DWTPs and less MP from WWTPs to receiving water environments.

Currently, polymeric membranes are more widely used in DWTPs/WWTPs than inorganic membranes due to their low cost and ease of production. Moreover, although the production and characterization of membranes and the application of membranes in treatment have been the most popular research topic for researchers in recent years, no study has been found that investigated MP removal by laboratory-scale membranes. In future studies, there is still a need to contribute to the literature by investigating the MP removal of produced polymeric and inorganic membranes. Since the characteristics of the membranes and the characteristics of the MPs affect the MP removal efficiency from water/wastewater, studies should be conducted to examine the MP removal efficiency with membranes from the water/wastewater after characterization of both membranes and MPs.

Currently, problems encountered in MP removal by membranes include contamination of membranes, the potential for MP release from polymeric membranes to water/wastewater, and the potential for MPs to pass even through dense membranes such as RO. There is a need to investigate which conditions cause the release of MPs from polymeric membranes to water/wastewater and cause MPs to pass through the membranes. In future studies, researchers should focus on producing membranes with antifouling and self-cleaning ability for MPs to minimize membrane fouling.

S.A., the sole author of the article, contributed 100% to the conceptualization, methodology, validation, research, resources, writing – original drafting, writing – reviewing and editing, and visualization.

All relevant data are included in the paper or its Supplementary Information.

The authors declare there is no conflict.