A review of removal of microplastics

Environment International 146 (2021) 106277 Contents lists available at ScienceDirect Environment International journal homepage: www.elsevier.com/locate/envint Review article A review of the removal of microplastics in global wastewater treatment plants: Characteristics and mechanisms Weiyi Liu a, Jinlan Zhang a, Hang Liu a, Xiaonan Guo a, Xiyue Zhang a, Xiaolong Yao b, Zhiguo Cao c, Tingting Zhang a, * a Department of Environmental Science and Engineering, Research Centre for Resource and Environment, Beijing University of Chemical Technology, Beijing 100029, People’s Republic of China Department of Environmental Science and Engineering, Beijing Technology and Business University, Beijing 100048, People’s Republic of China c School of Environment, Henan Normal University, Xinxiang 453007, People’s Republic of China b A R T I C L E I N F O A B S T R A C T Handling Editor: Frederic Coulon Wastewater treatment plants (WWTPs) are considered to be the main sources of microplastic contaminants in the aquatic environment, and an in-depth understanding of the behavior of microplastics among the critical treat ment technologies in WWTPs is urgently needed. In this paper, the characteristics and removal of microplastics in 38 WWTPs in 11 countries worldwide were reviewed. The abundance of microplastics in the influent, effluent, and sludge was compared. Then, based on existing data, the removal efficiency of microplastics in critical treatment technologies were compared by quantitative analysis. Particularly, detailed mechanisms of critical treatment technologies including primary settling treatment with flocculation, bioreactor system, advanced oxidation and membrane filtration were discussed. Thereafter, the abundance load and ecological hazard of the microplastics discharged from WWTPs into the aquatic and soil environments were summarized. The abundance of microplastics in the influent ranged from 0.28 particles L− 1 to 3.14 × 104 particles L− 1, while that in the effluent ranged from 0.01 particles L− 1 to 2.97 × 102 particles L− 1. The microplastic abundance in the sludge within the range of 4.40 × 103–2.40 × 105 particles kg− 1. In addition, there are still 5.00 × 105–1.39 × 1010 microplastic particles discharged into the aquatic environment each day Moreover, among the critical treatment technologies, the quantitative analysis revealed that filter-based treatment technologies exhibited the best microplastics removal efficiency. Fibers and microplastics with large particle sizes (0.5–5 mm) were easily separated by primary settling. Polyethene and small-particle size microplastics (<0.5 mm) were easily trapped by bacteria in the activated sludge of bioreactor system. The negative impact of microplastics from wastewater treatment plant was worthy of attention. Moreover, unknown transformation products of microplastics and their corresponding toxicity need in-depth research. Keywords: Microplastics Wastewater treatment technology Environmental toxicity Fate Meta-analysis 1. Introduction Microplastics widely occur in the atmosphere (Abbasi et al., 2019), soil (Guo et al., 2020), ocean (Wang et al., 2020b), freshwater (Han et al., 2020) and even in the sediment of an Arctic freshwater lake (González-Pleiter et al., 2020). They can adsorb pollutants, such as polycyclic aromatic hydrocarbons (Sørensen et al., 2020), heavy metals (Foshtomi et al., 2019), polybrominated diphenyl ethers (Singla et al., 2020), pharmaceutical and personal care products (Liu et al., 2019a; Ma et al., 2019c) from environmental media due to their small volume (particle debris size usually smaller than 5 mm) and high specific surface area (Thompson et al., 2004). As a result, microplastics always cause chronic toxicity due to their accumulation in organisms (Li et al., 2018). Wastewater treatment plants (WWTPs) are considered to be the main recipients of terrestrial microplastics before entering natural aquatic systems (Sun et al., 2019), which convert primary microplastics into secondary microplastics. The microplastics occurring in municipal wastewater commonly originate from daily human life activities. For example, polyester and polyamide components are commonly shed from clothing during the laundry process (Napper and Thompson, 2016), and personal care products such as toothpaste, cleanser and shower gel enter WWTPs resulting from our daily use (Magni et al., 2019). Moreover, the * Corresponding author. E-mail address: zhangtt@mail.buct.edu.cn (T. Zhang). https://doi.org/10.1016/j.envint.2020.106277 Received 27 July 2020; Received in revised form 6 November 2020; Accepted 7 November 2020 Available online 20 November 2020 0160-4120/© 2020 The Author(s). Published by Elsevier Ltd. This is an open (http://creativecommons.org/licenses/by-nc-nd/4.0/). access article under the CC BY-NC-ND license W. Liu et al. Environment International 146 (2021) 106277 plastics in garbage are decomposed by microorganisms in the leachate and then are discharged into WWTPs (Durenkamp et al., 2016). In addition, the microplastics floating in the atmosphere, which have been emitted by plastics industries and vehicles, also converge in WWTPs via atmospheric deposition (Liu et al., 2019c; Mintenig et al., 2017; Wright et al., 2020). It has been proven that untreated microplastics are commonly discharged from WWTPs, enter water bodies, and eventually accumulate in the environment (Carr et al., 2016). Therefore, it is urgent to study the performance of microplastic by different treatment tech nologies in WWTPs and understand the mechanism of removing microplastics to reduce the amount of microplastics entering the natural aquatic system. However, few pieces of research have been found to summarize the microplastics removal mechanisms of the critical treat ment technologies in the WWTPs. According to previous studies on the microplastic treatment tech nologies in WWTPs, microplastics were not completely removed from wastewater by these treatment technologies. For example, after the preliminary, primary, secondary and tertiary treatment processes in a WWTP in the UK, the overall abundance decreased by 6%, 68%, 92% and 96%, respectively (Blair et al., 2019). Mechanical, chemical, and biological treatment processes removed approximately 99% of the microplastics entering a WWTP (Ziajahromi et al., 2016). After treat ment, the removed microplastics were primarily transferred to the sludge phase (Ngo et al., 2019). Moreover, other noteworthy results, which are different from the results mentioned above, were obtained. For the same treatment process of microplastics, the microplastics removal efficiencies differed among various WWTPs. For example, aeration grit chambers, anaerobic-anoxicoxic (A2O) and advanced oxidation (UV and O3) processes were adopted as microplastic treatment methods in a Beijing WWTP, and their microplastics removal efficiencies were 58.84%, 54.47% and 71.67%, respectively (Yang et al., 2019). However, the microplastics removal efficiencies for the same treatment processes in a Shanghai WWTP decreased to 49.56%, 26.01%, and 0.78%, respectively (Jia et al., 2019). These results indicate that it is very challenging to understand the role of a given treatment process in microplastics removal in a WWTP via a single investigation. Moreover, the conventional study methods on microplastics removal are mainly based on qualitative analysis rather than quantitative analysis (Ngo et al., 2019). Therefore, it is necessary to develop new methods to qualitatively assess the removal performance of microplastics in WWTPs. In recent years, as a statistical method for the quantitative analysis of a series of independent features of the same object, meta-analysis has been increasingly applied to analyze wastewater problems in a more scientific manner (Erni-Cassola et al., 2019). For example, meta-analysis results indicated that photocatalysts generally attain the highest diaz inon elimination efficiency with an average efficiency of 79.2% (95% confidence interval: 76.8%–81.5%) (Malakootian et al., 2020). Another meta-analysis study revealed that membrane bioreactor systems might present the highest removal efficiency of organic trace contaminants in wastewater (Melvin and Leusch, 2016). To date, no qualitative assess ment of the removal of microplastics in WWTPs via meta-analysis has been reported. It is believed that the meta-analysis approach can provide a better understanding of the characteristics of microplastics in WWTPs and a more accurate estimate of microplastics removal in critical wastewater treatment technologies. In this study, the critical microplastic treatment technologies in global WWTPs are reviewed. Meta-analysis was first performed to identify the optimal microplastics removal technology. Thirty-eight WWTPs in eleven countries worldwide were investigated in terms of the occurrence, transfer, and removal mechanism of microplastics in different critical treatment technologies. The discussion focused on the removal behavior of various microplastic shapes, polymer types, and particle sizes. In addition, the risks of microplastics to the aquatic and soil environments were also evaluated. The results are instructive for a scientific understanding of the fate of microplastics in WWTPs. 2. Methods 2.1. Data collection The publications were obtained by searching all databases in the China National Knowledge Infrastructure and Web of Science using search terms such as microplastic, wastewater treatment plant, sewage treatment plant, and sludge. The search included all available publica tions until April 2020. The search results indicated that microplastic was first defined in Science by Thompson in 2004 (Thompson et al., 2004). The first study on the occurrence and removal of microplastics in WWTPs was published in 2015 by Talvitie et al. (2015). Relevant pub lications over the past three years (2018–2020) have been reviewed. The publications were individually assessed to eliminate irrelevant articles based on their abstracts, tables, and figures. Ultimately 23 highly rele vant papers covering microplastics in global WWTPs were considered for meta-analysis. Data retention criteria in publications included: 1) the name of the corresponding treatment process and microplastic abun dances of the influent and effluent; 2) polymer types, shapes and particle sizes of microplastics and their abundances; 3) studies are independent of each other and there are no duplicate studies; and 4) basic informa tion of WWTPs (location, daily capacity, serving population, etc.). GetData Graph Digitizer (v.2.25) was used to extract the data of the microplastic abundances (the influent and effluent of treatment tech nology) and removal efficiency presented in the graph. Due to the difficulty in measuring the mass of microplastics, the behavior from the perspective of the quantity abundance (particle L− 1) was evaluated in this study. The microplastic abundance and removal efficiency were generally provided in the publications. Otherwise, nondetection was assigned a zero value (Yang et al., 2018). It was clear that the inconsistent microplastic shape nomenclatures and size classes were discussed in the existing research publications. An ambiguous nomen clature inhibits research progress, leading to confusion and miscom munication (Hartmann et al., 2019). For analysis convenience, the film or sheet, pellet or spherical or bead and particle or granule were consistently renamed film, pellet and particle, respectively. The micro plastic size classifications were 100 μm, 500 μm, 1 mm and 5 mm in previous studies (Wang et al., 2019). Thus, this study divided the size in the meta-analysis as smaller than 0.5 mm, 0.5–1 mm and 1–5 mm. 2.2. Quantitative meta-analysis Meta-analysis was performed in R-project for Statistical Computing (version 3.6.2) using the meta package (Schwarzer, 2007). The micro plastic data obtained from each independent study was analyzed, and the microplastics removal of different treatment technologies was compared. The risk ratio (RR) value was assigned as the effect size in a single study. The calculation method of the RR value can be found elsewhere (Schwarzer, 2007). An RR value lower than 1 indicated that microplastics could be removed, and lower the value is, the better the removal is. The effect size in the meta-analysis was a weighted average of each single study. A study with a higher weight value was supposed to impose a greater influence on effect size. The weights were determined by the random effects model of meta-analysis, due to the high hetero geneity among the studies (McKenzie et al., 2016). RR = AI AE (1) where AI and AE denote the event probabilities for the experimental and control groups, respectively. In this study, meta-analysis was used to investigate the removal ef ficiency of microplastics in the primary, secondary, and tertiary treat ment processes, and also to compare the removal efficiency between critical treatment technologies. Further, we also analyzed the removal efficiency of different classifications of microplastics (shapes and 2 W. Liu et al. Environment International 146 (2021) 106277 particle sizes) in these treatment processes. τ2 = 2.3. Heterogeneity and publication bias W= The heterogeneity test aims to determine whether genuine differ ences exist between study results (Higgins et al., 2002). The heteroge neity can be expressed by I2, τ2, or the Cochran Q test (Higgins et al., 2003; Langan et al., 2019). The I2 quantity, ranging from 0% to 100%, describes the degree of inconsistency among studies in a meta-analysis sense. The larger the I2 quantity is, the larger the difference between studies is. The heterogeneity variance parameter is denoted as τ2, which effectively reflects the heterogeneity among studies. The Cochran Q test statistic is computed by summing the squared deviations of each study’s estimate from the overall meta-analysis estimate, thereby weighting each study’s contribution in the same manner as conducted in the metaanalysis. The p-value is obtained by evaluating the chi-square distribu tion with (k-1) degrees of freedom (df) (k: the number of studies). The difference among studies is caused by random errors when the p-value is smaller than 0.05. ∑ Q= ωi (θi − θ)2 (2) I2 = Q − df × 100% Q I2 W 1 − I2 df Σωi ∑ (Σωi )2 − (4) (5) ωi 2 where ωi is the weight of each study, θi is the effect value (RR) of each study, θ is the average of the effect value (RR), and df is the degree of freedom (k-1). Contour enhanced funnel plots were used to test for publication bias. The funnel plots were performed in R-project using the meta package. These plots showed effect sizes and standard errors in each metaanalysis. The effect sizes which were symmetrical and on the top of the funnel proved there was no bias (Egger et al., 1997). 3. Characteristics of the microplastics in WWTPs 3.1. Microplastics in wastewater Microplastics were frequently detected in the influent and effluent of WWTPs. Table 1 lists the location, daily treatment capacity, serving population, source of wastewater and main treatment technologies of the WWTPs in this study. The microplastic abundances in the primary, secondary, and tertiary treatment processes and effluent are (3) Table 1 Information of WWTPs in this study. Location Capacity (m3/day) Population Treatment processes Source of wastewater Reference R1 R2 R3 R4(W1) R4(W2) R5(W1) R5(W2) R5(W3) R6 R7(W1) Scotland, UK Cartagena, Spain Madrid, Spain Hong Kong, China Hong Kong, China Daegu, Korea Daegu, Korea Daegu, Korea Wuhan, China M-City, Korea 166,422 35,000 28,400 93,000 2,400,000 26,545 469,249 20,840 20,000 – 1.8 × 105 2.1 × 105 – – – – – – – – Pri, Sec, Ter (Nitrification) Pri, Sec Sec (A2O) Pri, Sec Pri, Ter (Chlorination) Pri, Sec, Ter (Coagulation, O3) Pri, Sec, Ter (Coagulation, DF) Pri, Sec, Ter (Coagulation, RSF) Pri, Sec (A2O), Ter (Chlorination) Pri, Sec (A2O) Municipal Municipal and Industrial Municipal Municipal After primary treatment Municipal and Industrial Municipal and Industrial Municipal and Industrial Industrial, Agricultural, Municipal Municipal Blair et al., 2019 Bayo et al., 2020 Edo et al., 2020 Ruan et al., 2019 Ruan et al., 2019 Hidayaturrahman and Lee, 2019 Hidayaturrahman and Lee, 2019 Hidayaturrahman and Lee, 2019 Liu et al., 2019d Lee and Kim, 2018 R7(W2) R7(W3) R8 R9(W1) R9(W2) R10(W1) R10(W2) R11(W1) R11(W2) R11(W3) R11(W4) R12 R13(W1) R13(W2) Y-City, Korea S-City, Korea Mikkeli, Finland Sydney, Australia Sydney, Australia Wuxi, China Wuxi, China Helsinki, Finland Turku, Finland Hameenlinna, Finland Mikkeli, Finland Scotland, UK Shanghai, China Shanghai, China – – 10,000 17,000 48,000 50,000 70,000 – – – – 260,954 – – – – – 6.7 × 1.5 × – – – – – – 6.5 × 3.5 × 2.9 × Sec (SBR) Pri, Sec Pri Sec Sec Pri, Sec (OD), Ter (UV) Pri, Sec (A2O + MBR) Ter (DF) Ter (RSF) Ter (DAF) Ter (MBR) Pri, Sec Pri, Sec (A2O), Ter (UV) Pri, Sec (A/O), Ter (UV) Municipal Municipal Municipal Municipal Municipal Municipal Municipal Municipal Municipal Municipal Municipal Municipal Municipal Municipal Lee and Kim, 2018 Lee and Kim, 2018 Lares et al., 2018 Ziajahromi et al., 2017 Ziajahromi et al., 2017 Lv et al., 2019 Lv et al., 2019 Talvitie et al., 2017a Talvitie et al., 2017a Talvitie et al., 2017a Talvitie et al., 2017a Murphy et al., 2016 Jia et al., 2019 Jia et al., 2019 R14 R15 R16 R17 R18 R19(W1) R19(W2) R20 R21(W1) R21(W2) R22 R23(W1) Helsinki, Finland Beijing, China Vancouver, Canada Helsinki, Finland Northern, Italy Detroit, USA Detroit, USA Paris, France Xiamen, China Xiamen, China Xiamen, China Los Angeles, USA 270,000 1,000,000 493,271 – 400,000 2,500,000 1700 240,000 75,000 245,800 257,936 – 8.0 × 2.4 × 1.3 × 8.0 × 1.2 × – – – 3.4 × 1.2 × 1.0 × – Pri, Sec, Ter (BAF) Pri, Sec (A2O), Ter (UF, UV, O3) Pri, Sec Pri, Sec, Ter (BAF) Pri, Sec, Ter (SAF) Pri, Sec, Ter (Chlorination) Pri, Sec, Ter (Chlorination) Pri, Sec (Biofilter) Pri, Sec Pri, Sec Pri, Sec (Biofilter) Ter (GF) Municipal Municipal Municipal Municipal Combined sewers Raw wastewater and stormwater Raw wastewater and stormwater Municipal and Industrial Municipal and Industrial Municipal and Industrial Municipal and Industrial – Talvitie et al., 2017b Yang et al., 2019 Gies et al., 2018 Talvitie et al., 2015 Magni et al., 2019 Michielssen et al., 2016 Michielssen et al., 2016 Dris et al., 2015 Wang et al., 2019 Wang et al., 2019 Long et al., 2019 Carr et al., 2016 R23(W2) R24 Los Angeles, USA Oldenburg, Germany – 35,616 – 2.1 × 105 Ter (Centrata thickening) Ter (PF) – Municipal and Industrial Carr et al., 2016 Mintenig et al., 2017 104 105 105 106 106 105 106 106 105 106 105 106 106 (1) Pri, Sec, Ter refer to primary treatment, secondary treatment and tertiary treatment (2) A2O: anaerobic-anoxic-oxic; A/O: anoxic oxic; OD: oxidation ditch; DF: disc filter; RSF: rapid (gravity) sand filter; DAF: dissolved air flotation; BAF: biologically active filter; GF: gravity filter; PF: post-filtration; SAF: sand filter; UF: ultra-filtration 3 W. Liu et al. Environment International 146 (2021) 106277 summarized in Table S1. The microplastic abundances in the influent of the WWTPs ranged from 0.28 particles L− 1 to 3.14 × 104 particles L− 1 (mean value: 1.90 × 103 particles L− 1; median value: 57.60 particles L− 1). The differences in the microplastic abundance could be related to a variety of complex factors, such as the population served, wastewater sources (municipal or industrial), economy, and lifestyle. In the municipal WWTPs, the microplastic abundance was lower, ranging from 0.28 particles L− 1 to 6.10 × 102 particles L− 1 (mean value: 1.27 × 102 particles L− 1; median value: 31.10 particles L− 1). In the municipal and industrial WWTPs, the microplastic abundances ranged from 1.60 par ticles L− 1 to 3.14 × 104 particles L− 1 (mean value: 5.23 × 103 particles L− 1; median value: 1.86 × 102 particles L− 1). However, few studies have investigated the microplastic emissions directly from plastic processing industrial WWTP. In terms of the serving population, the abundance of the influent microplastics was positively correlated with the serving population in most WWTPs (Mason et al., 2016). The microplastic abundance was also influenced by the sampling and detection methods. Limited wastewater sample volumes increased the uncertainty of the microplastic abundance and experimental errors. These problems enhanced the difficulties in the research on the microplastic fate in the WWTPs. As indicated in Table S1, microplastics were detected in all treatment processes in the WWTPs. The microplastic abundance gradually decreased from primary treatment to secondary treatment. The primary treatment process based on physical mechanisms was considered the first barrier to remove microplastics in WWTPs. Primary settling tank was the most commonly implemented primary treatment method. The microplastic abundance after primary treatment processes ranged from 0.22 particles L− 1 to 1.26 × 104 particles L− 1 (mean value: 6.87 × 102 particles L− 1; median value: 4.90 particles L− 1). Their abundance decreased by 4.06–98.96% (mean value: 56.75%; median value: 54.88%), compared to the abundance of microplastics in the influent. After the primary treatment process, biological treatment (secondary treatment process) was the most critical technology in the WWTPs. In biological treatment, A2O was the most widely used technology in WWTPs. Meanwhile, the biofilter technology had a high biomass load and a high volumetric reaction rates, which improved the pollutant removal efficiency and gradually became more popular in WWTPs (Liu et al., 2020a; Rocher et al., 2012). The abundances after secondary treatment processes ranged from non-detection to 7.86 × 103 particles L− 1(mean value: 4.67 × 102 particles L− 1; median value: 6.90 particles L− 1), resulting in a decrease of abundance by 20.45–95.45% (mean value: 66.63%; median value: 73.53%). To further remove the contaminants, 61.76% of the WWTPs employed tertiary treatment processes, such as advanced oxidation and membrane filtration processes. After tertiary treatment processes, the microplastic abundance further decreased in most of the investigated WWTPs (85.71%), while in others, the abundance increased. The abundance of the microplastics in effluent ranged from non-detection to 2.97 × 102 particles L− 1 (mean value: 19.26 particles L− 1; median value: 0.40 particles L− 1). Compared with the influent, the microplastic abundance decreased by 50.00–99.57% (mean value: 85.58%; median value: 90.34%). As a result, at most 50.00% of the microplastics in the WWTPs was still discharged through the effluent and entered the receiving water systems. Hence, microplastic-targeted treatment pro cesses are urgently needed. the WWTPs. 3.2.1. Shape The shape is an important classification factor of microplastics. The shape of microplastics affects their removal efficiency in WWTPs (McCormick et al., 2014). Nine shapes of microplastics were detected in the influent and effluent of the WWTPs. The abundances of the different microplastic shapes observed in the WWTPs are summarized in Table 2. Fibers, pellets, fragments, and films were the most widely detected microplastics in wastewater, and their highest abundances were 91.32%, 70.38%, 65.43%, and 21.36%, respectively (Bayo et al., 2020; Hidayaturrahman and Lee, 2019; Lares et al., 2018). The fiber, a filamentary microstructure, was the dominant micro plastic shape in the WWTPs. The microplastic fibers originated from domestic washings. The increasing amount of washing and textile con sumption resulted in the more frequent detection of fibers (Cesa et al., 2017). During the textile production process, fibers are also produced in shearing and splicing processes, after which they enter wastewater (Hidayaturrahman and Lee, 2019; Napper and Thompson, 2016). The microplastic fragments and pellets originated from cosmetics and per sonal care products, such as toothpaste, masks, and soaps (Carr et al., 2016). The microplastic films originated from plastic packing bags (Kazour et al., 2019). Moreover, other microplastic shapes, such as foams, particles, ellipses, lines, and flakes, were also detected in the WWTPs. 3.2.2. Particle size Microplastics may end up in the food chain, and the size of micro plastics rather than their shape was a crucial factor influencing their performance and transformation in the WWTPs (Lehtiniemi et al., 2018). Therefore, it is important to highlight the particle size of microplastics. The distribution of the microplastic particle sizes in the WWTPs is shown in Fig. S1. The abundance of the microplastics smaller than 1 mm ranged 65.0–86.9% in the influent and 81.0–91.0% in the effluent. With decreasing microplastic sizes, the primary microplastics were crushed (physical, chemical, and biological processes) into secondary micro plastics (Magni et al., 2019). The smaller microplastic particles were more likely to be ingested by plankton, filter-feeding organisms, and fishes, which can cause a series of toxicological effects in these organ isms (Qiao et al., 2019). Therefore, the research of the particle size of microplastics, especially the smaller particle size (smaller than 1 mm) can be of guiding significance for the subsequent study of biological toxicity and the environmental transformation of microplastics. 3.2.3. Polymer type The abundances of the different microplastic polymer types in the influent and effluent are listed in Table 3. Twenty-nine kinds of poly mers were detected in the influent and effluent of the WWTPs. Poly ethene (PE), polypropylene (PP), polyamide (PA), polyester (PES), polystyrene (PS) and polyethene terephthalate (PET) were the top six most widely detected microplastics in the wastewater, and their highest Table 2 The abundance of different shapes of microplastics in WWTPs. Shape Fiber Fragment Film Pellet Foam Particle Ellipse Line Flake 3.2. Shape, particle size and polymer type distribution in the influent and effluent Microplastics are a type of polymer mixture with various shapes and sizes. Different shapes and sizes of microplastics possessed different physicochemical and toxicity properties (Lehtiniemi et al., 2018). Therefore, this study emphasized the occurrence and removal of microplastics with different shapes, particle sizes, and polymer types in Influent (particles L− 1) 3 0.22–4.60 × 10 0.25–3.40 × 103 0.06–1.30 × 103 0.01–2.21 × 104 nd-2.33 nd-2.91 × 102 0.36 0.12 0.92 Note: nd means non-detection. 4 Effluent (particles L− 1) Detection times nd-35.00 nd-80.00 nd-12.00 nd-1.33 × 103 nd nd-10.00 nd 0.12 nd 12 11 9 7 4 3 1 1 1 W. Liu et al. Environment International 146 (2021) 106277 (Sun et al., 2019). Table 4 summarizes the microplastic abundance levels in the sludge treated by different treatment technologies within the range from 4.40 × 103 particles kg− 1 to 2.40 × 105 particles kg− 1. The microplastic abundance in the sludge from the primary treatment pro cess was higher than that of the secondary treatment process. Gies et al. (2018) estimated that 0.54–1.28 trillion microplastics occurred in pri mary sludge and 0.22–0.36 trillion microplastics occurred in secondary sludge. In addition, Talvitie et al. (2017b) calculated that 20% of the microplastics in the secondary sludge flowed back into the wastewater. Sludge utilization has received much attention in recent years. The sludge from the WWTPs was mainly utilized for agricultural purposes in Norway (82%), Ireland (63%), the US (55%), China (45%) and Sweden (36%), and it was incinerated in the Netherlands (99%), Korea (55%) and Canada (47%), while the sludge was applied as soil compost in Finland (89%) and Scotland (40%) (Fig. 1) (Rolsky et al., 2020). Soil contaminated with microplastics represents a great threat to crops and agricultural products. Corradini et al. (2019) reported that the average microplastic abundance in agricultural soils originating from sludge disposal was 3,500 particles kg− 1. Pyroplastics are a new type of pollutant derived from the informal or more organized burning of manufactured microplastics. After sludge incineration, pyroplastics enter the environment and cause great threats (Turner et al., 2019). In China, 35% of the sludge originating from WWTPs still enter landfills (Fig. 1). Microplastics are further transferred into the soil and ground water through the leachate (Chen et al., 2012; Rolsky et al., 2020). In general, soil contamination of microplastic is scarcely known and is thus considered one of the pressing concerns related to microplastics. Table 3 The abundance of different polymer types of microplastics in WWTPs. Polymer Abbreviation Influent (particles L− 1) Effluent (particles L− 1 ) Detection times Polyethene Polypropylene Polyamide Polyester Polystyrene Polyethene terephthalate Polyurethane Polyvinyl chloride Polyvinyl acetate Alkyd Ethylenevinylacetate Polyacrylates Acrylic Polyvinyl ethylene Polyvinyl fluoride Styrenebutadienestyrene Styrene-ethylenebutadieneStyrene Styrene acrylonitrile Polyvinyl alcohol Acrylamide Polyethene& Polypropylene Paint Polystyrene acrylic Polyvinyl acrylate Styrenevinyltoluenebutylacrylate Polyterpene Acrylonitrilebutadiene Ehtylene-acrylate Ehtylenepropylene PE PP PA PES PS PET 0.03–1.05 0.02–1.42 0.06–0.71 0.22–6.31 0.00–0.41 0.01–0.63 0.00–0.67 0.00–0.22 0.00–0.06 0.07–1.33 0.00–0.08 0.00–0.16 9 8 6 6 5 5 PUR / PU PVC PVA – EVA 0.07–1.40 0.12–1.65 0.26–0.50 0.13–4.51 0.00–0.01 0.00–0.02 0.00 0.00–0.01 0.00–0.02 0.00 4 3 2 2 2 – – PVE PVF SBS 0.06–0.40 1.30 0.09 0.09 0.02 0.00–0.03 0.03 0.00 0.00 0.00 2 1 1 1 1 SEBS 0.06 0.00 1 SAN 0.01 0.00 1 PVAL – PE&PP 0.03 0.09 0.09 0.00 0.00 0.01 1 1 1 – PS acrylic PV acrylate – 0.01 0.30 0.09 0.01 0.00 0.00 0.00 0.00 1 1 1 1 – – 0.03 0.80 0.01 0.01 1 1 – – 0.14 0.28 0.01 0.00 1 1 4. Removal of microplastics in the WWTPs 4.1. Comparison of the different treatment technologies for microplastics removal Previous studies reported the microplastics removal efficiency via individual field sample collected from primary, secondary, and tertiary treatment processes. However, these studies could not accurately determine the optimal treatment process and mechanism of micro plastics removal. Therefore, this study compared the removal efficiency of different treatment technologies in global WWTPs via meta-analysis. The different treatment technologies applied in the WWTPs included primary treatment processes (primary settling treatment, grit and grease treatment), secondary treatment processes (A2O, biofilters, and other bioreactors) and tertiary treatment processes (UV, O3, chlorination, biologically active filters (BAFs), disc filters (DFs), and rapid sand filters (RSFs)). Fig. 2 shows the meta-analysis results for the different treatment technologies. The weighted average RR values of the primary, second ary, and tertiary treatment processes were 0.40, 0.39, and 0.48, respectively (Fig. S2). The primary and secondary treatment processes attained similar microplastics removal efficiencies. The tertiary treat ment process achieved limited removal efficiency. Especially, the abundances were 64.07%, 32.92%, 10.34%, 75.36%, 24.17%, and 28.90%, respectively (Long et al., 2019; Mintenig et al., 2017; Talvitie et al., 2017a; Ziajahromi et al., 2017). The PE, PP, and PS microplastics originated from plastic products, including food packaging bags, plastic bottles, and plastic cutlery (Lares et al., 2018; Mintenig et al., 2017; Talvitie et al., 2017b). The PA, PET, and PES microplastics mainly originated from textiles and synthetic clothing, which are the main sources of household microplastics (Hernandez et al., 2017; Sun et al., 2019; Wei et al., 2019). Furthermore, the mechanical crushing of plastic products, the tire, and textile industries and the rubber particles in road dust were also identified as potentially important sources of the PE, PP, PS and PES microplastics (Hidayaturrahman and Lee, 2019; Nizzetto et al., 2016; Talvitie et al., 2017a). In addition to the polymer types mentioned above, specific polymers were also identified in the WWTPs. For example, alkyds, which are widely used in industrial coatings, exhibited the highest abundance in a Glasgow WWTP (28.67%) (Murphy et al., 2016). Therefore, research priority should be assigned to specific polymer types in addition to common polymers. Table 4 The abundance of microplastics in the sludge of different wastewater treatment processes. R6 R7 (a) R7 (b) R7 (c) R16 R16 3.3. Microplastics in the sludge Most of the microplastics removed from wastewater were retained in the sludge (Mahon et al., 2017). It was found that the microplastic abundance in the sludge was much higher than that in the wastewater 5 Treatment Process Abundance (Particles kg− 1) Primary clarifier + A2O + Secondary clarifier Primary settling tank + A2O + Secondary settling tank SBR 2.40 × 105 1.49 × 104 9.65 × 103 Primary settling tank + Secondary settling tank 1.32 × 104 Primary settling Secondary clarifiers 1.49 × 104 4.40 × 103 W. Liu et al. Environment International 146 (2021) 106277 poor efficiency in microplastics removal. But a recent study showed that 69–79% of microplastics entering WWTPs are removed by screening and grit treatment (Ziajahromi et al., 2021). Only light floating microplastics could be removed during the grease skimming process (Sun et al., 2019). Grit and grease process combined with primary settle treatment could improve the efficiency of removing microplastics. Bioreactor (except A2O and biofilter) attained a notable microplastics removal efficiency. Sequence batch reactor process (SBR) was found to have removal 98% microplastics (Lee and Kim, 2018). However, A2O technology was not suitable for the removal of microplastics because of the sludge returned. Jiang et al. (2020) indicated that the anoxic-oxic process (A/O) captured about 16.9% of microplastics in wastewater. The removal efficiency of the same treatment process is closely related to the characteristics of wastewater and the types of microplastic polymers. Advanced oxidation processes showed the medium removal efficiency of microplastics. It can be seen that the removal efficiencies of microplastics in critical treat ment technologies are quite different. 4.2. Influence of the microplastic shape, particle size and polymer type on microplastics removal Fig. 3 shows the meta-analysis results of microplastics removal for four shapes by different treatment technologies which was calculated with the data summarized in Section 3.2.1. Among the four microplastic shapes, fibers were the most widely detected microplastics in waste water. Their weighted average RR values in the primary, secondary and tertiary treatment processes were 0.31, 0.41, and 0.43, respectively (Fig. S3). Primary treatment has superiority over secondary and tertiary treatment for fiber microplastics removal. Fibers were easily trapped during primary treatment due to flocculation and settling. After the primary treatment process, most of the easily settled or skimmed par ticles were removed, but the remainder might exhibit a neutral buoy ancy (Sun et al., 2019). In contrast, fragments exhibited excellent removal efficiency during the secondary treatment process. The weighted average RR values of fragments were 0.41, 0.30, and 0.36, respectively (Fig. S3). Fragments with a lamellar structure gradually agglomerated and were ingested by the activated sludge (Jeong et al., 2016). The weighted average RR values of the films in the primary, secondary, and tertiary treatment processes were 0.35, 0.34, and 0.47, respectively (Fig. S3). The microplastics removal efficiencies of the primary and secondary treatment processes were higher than that of the tertiary treatment processes. Pellets were easier to remove during the tertiary treatment process. The weighted average RR values of the pel lets in the primary, secondary and tertiary treatment processes were 0.63, 0.76, and 0.35, respectively (Fig. S3). Both filter-based and Fig. 1. Proportions of sludge utilization type in 12 countries (the remaining utilization types were described as ‘others’). Fig. 2. Meta-analysis results of microplastics removal by different treat ment processes. tertiary treatment processes exhibited a wide range of the 95% CIs (0.22–1.06) because of the difference between the advanced oxidation process and filter technology. Among them, the advanced oxidation treatment process removed pollutants via chemical methods, while filter technology removed pollutants through physical methods. The micro plastics removal efficiency of the critical treatment technologies fol lowed the sequence of biofilters, filters, primary settling, bioreactors (except for A2O and biofilters), grit and grease removal with primary settling, advanced oxidation, grit and grease removal, and A2O, with weighted average RR values of 0.32, 0.33, 0.39, 0.41, 0.42, 0.56, 0.61 and 0.73, respectively (Fig. S2). Therefore, filter-based technologies (biofilter, ultrafiltration (UF), rapid sand filter (RSFs), etc.) achieved the best performance in removing microplastics. Among them, RSF technology provides rapid and efficient removal of microplastics (Talvitie et al., 2017a). But in this process, the microplastics will be broken into smaller particles (Prata, 2018; Sol et al., 2020). Primary settling treatment attained an excellent efficiency in removing microplastics, while grit and grease treatment exhibited a Fig. 3. Meta-analysis results of microplastics removal with different shapes by different treatment processes. 6 W. Liu et al. Environment International 146 (2021) 106277 promising. The impact of different shapes, particle sizes, and polymer types on microplastics removal in different treatment processes should receive more attention. What’s more, the mechanisms of microplastics removal by the critical treatment technologies should also be studied in depth. advanced oxidation treatment processes could effectively intercept pellets. Moreover, the tertiary treatment process also presented an extremely high efficiency in removing microplastics with specific properties and very small particle sizes. Fig. 4 shows the meta-analysis results of microplastics removal for three particle sizes by the different treatment processes. During the tertiary treatment process, the microplastics usually had a small particle size (smaller than 0.5 mm), which was more difficult to detect and remove (Hidalgo-Ruz et al., 2012). In addition, in the advanced oxida tion processes, microplastics were continuously crushed, resulting in a negative removal (Lv et al., 2019; Ruan et al., 2019). Therefore, the meta-analysis was only conducted on the primary and secondary treat ment processes. Microplastics with particle sizes smaller than 0.5 mm were easily trapped during secondary treatment processes. The weighted average RR values in the primary and secondary treatment processes were 0.70 and 0.48, respectively (Fig. S4). Microplastics with a particle size ranging from 0.5 mm to 1 mm were better removed during primary treatment processes, as well as microplastics with a particle size ranging from 1 mm to 5 mm. The weighted average RR values of the microplastics in the 0.5–1 mm size range were 0.31 and 0.74, respec tively (Fig. S4). The weighted average RR values of the microplastics in the 1–5 mm size range were 0.06 and 0.53, respectively (Fig. S4). On the one hand, the fibers and films had a large particle size and low density, and they were easily removed by flotation and grease removal processes. On the other hand, the pellets in personal care products had high den sity, and they generally sank to the bottom of pools due to gravity (Lehtiniemi et al., 2018). Common polymer analytical methods included gas chromatography coupled to mass spectrometry (Dümichen et al., 2017), liquid chroma tography (Elert et al., 2017), Fourier transform infrared spectroscopy (Mintenig et al., 2017) and Raman spectroscopy (Lares et al., 2018). However, it was still difficult to identify individual microplastic polymer type due to the limitation of these analytical methods (Hidalgo-Ruz et al., 2012). Thus, a meta-analysis for the different polymer types could not be conducted until now. PE, as the most frequently detected microplastic polymer type in the WWTPs, was efficiently removed during the secondary treatment process, which was also true for PS polymer. The positively charged PE and PS microplastics had a high affinity for the negatively charged activated sludge mass (Bhattacharya et al., 2010). In a word, fibers and microplastics with large particle sizes (0.5–5 mm) were easily separated by primary settling. Polyethene (PE) and small-particle size microplastics (smaller than 0.5 mm) were easily trapped in the activated sludge by bacteria. Therefore, choosing suitable treatment technologies for microplastics removal in WWTPs is quite 4.3. Mass of microplastics The quantification of microplastics focused on determining the number of particles, because the influence and behavior of microplastics is closely related to the number of particles (Andrady, 2011). Because of the aging of the microplastics in the environment, the microplastics were broken into small particles. Therefore, using the mass of micro plastics to supplement the description of the occurrence of microplastics could scientifically and accurately quantify the load of microplastics in the environment (Rocha-Santos and Duarte, 2015). Mass balance was used as an intuitional way to explore the fate of microplastics in WWTPs. From a mass point of view, the WWTP has shown excellent perfor mance in the removal of microplastics. Simon et al. (2018) investigated the mass of microplastics in the influent and effluent of 10 WWTPs in Denmark for the first time. The mass of microplastics in the influent ranged from 61 μg L− 1 to 1189 μg L− 1 (mean value ± standard deviation: 352 ± 324 μg L− 1; median value: 240 μg L− 1). The mass of microplastics in the effluent ranged from 0.5 μg L− 1 to 11.9 μg L− 1 (mean value ± standard deviation: 4.4 ± 4.3 μg L− 1; median value: 3.7 μg L− 1). Through the mass balance of the microplastics in the WWTP, it can be found that only 0.22–6.23% of the microplastics will enter the natural aquatic system through the effluent. Furthermore, Lv et al. (2019) systematically evaluated the mass of microplastics in the effluent of each critical treatment process from two WWTPs in Wuxi, Jiangsu Province, China. The mass of the microplastics in the influent of two WWTPs was 280 ± 4.5 kg per day and 392 ± 4.5 kg per day, respectively. After the primary treatment process (aerated grit chamber/ rotary grit chamber), the mass of the microplastics was reduced to 271.6 ± 2.5 kg and 388 ± 2.5 kg per day, respectively. After the secondary treatment process (oxidation ditch/ anaerobic-anoxic-oxic), the mass of the microplastics was reduced to 225 ± 5.0 kg and 329 ± 3.5 kg per day, respectively. After the tertiary treatment process (UV disinfection/ Membrane tank), the mass of microplastics in the effluent was 8.4 kg and 1.96 kg per day, respectively. Mass of microplastics accumulated in excess sludge was 0.51 kg per day from oxidation ditch system and 0.033 kg per day from anaerobic-anoxic-oxic-membrane tank system. It can be seen that only 3% or less of the microplastics in the wastewater treatment plant were discharged into the aquatic environment and 1% of the microplastics were retained in the sludge produced by the biochemical treatment process. The remaining 96% of the microplastics were degraded, skim ming or retained by the membrane. Among them, membrane retention occupied a larger proportion, which can be seen from the high removal rate of the filter-based process. 4.4. Publication bias Contour enhanced funnel plots of the microplastics removal by pri mary, secondary, and tertiary treatment processes were presented in Fig. S5. Contour enhanced funnel plots of the different shapes and par ticle sizes microplastics removal by different treatment processes were presented in Fig. S6 and Fig. S7, respectively. Collectively, although the extractable studies are relatively few in some of the meta-analysis, no evidence of publication bias was observed in the funnel plots. 4.5. Mechanisms of critical treatment technologies in microplastics removal 4.5.1. Primary settling with flocculation Fig. 5 showed the schematics of flocculation and primary settling technologies in microplastics removal. In the flocculation process, flocs Fig. 4. Meta-analysis results of microplastics removal with different particle sizes by different treatment processes. 7 W. Liu et al. Environment International 146 (2021) 106277 Fig. 5. The schematics of primary settling with flocculation technologies in microplastics removal (Lapointe et al., 2020). Fig. 6. The schematics of the bioreactor system in microplastics removal. (A) Activated sludge process (Zhang et al., 2020); (B) MBR (Adapted from Li et al., 2020); (C) Biofilter; (D) A2O (Liu et al., 2020b). 8 W. Liu et al. Environment International 146 (2021) 106277 interacted with microplastics via hydrogen bonds, van der Waals forces, or electrostatic forces (Duan and Gregory, 2003; Lapointe et al., 2020). Like-charged microplastic particles maintain stability due to the repul sive inter-particle electrostatic forces. Flocculants possessing opposite charges with microplastics effectively reduced the repulsive potential between microplastic particles. It was possible for the Brownian motion and mechanical agitation to become operative, leading to microplastics and flocs aggregation (Larue et al., 2003). Both iron-based salts and aluminum-based salts are widely used flocculants in wastewater treatment (Ma et al., 2019a). The flocculation of microplastics with iron was caused by the adsorption of iron hy droxide aggregates (Larue et al., 2003). Small aggregates with high positive charge were locally adsorbed on the microplastic surfaces at low pH media. In this case, flocs neutralized microplastic charges and eliminated repulsion forces between microplastics. At neutral and basic pH media, the floc aggregates size increased, and aggregated form bridges between microplastics (Larue et al., 2003; Ma et al., 2019b). Microplastics interacted with aluminum-based flocculants via hydrogen bonds. Cationic aluminum flocculant was also electrostatically attached to anionic carboxyl groups in the weathered microplastics. The presence of new functional groups (such as hydroxyl group (OH), carboxyl group – C)) on the weathered (COOH), and carbon-carbon double bond (C– microplastic surface promoted interactions between the flocs and microplastics (Lapointe et al., 2020). Subsequently, the primary settling technology mainly removed the settable parts in the suspended microplastics. Most of the non-sinkable floating microplastics adhered to the flocs and precipitated together, others were skimmed as scum (Lee et al., 2012). These microplastics were discharged as primary sludge (Murphy et al., 2016). Existing studies proposed the mechanisms for the removal of microplastics in flocculation technology and primary settling technology but lack of identification of the degradation products of microplastics and their physiochemical properties. The toxicity of the substances generated after the flocculation and the impact of settling time on settling effi ciency of primary settling technology are less known. microplastics removal in the MBR system. What’s more, the removal of microplastics could be related to the size of microplastics. The mem brane applied in the MBR system usually has a pore size of 0.1 μm. Thus, the microplastics could be removed in the MBR system theoretically. (Li et al., 2020). The biofilter technology was applied as a deep treatment unit after the bioreactor system. The microplastics entered the biofilter treatment unit have smaller particle size and lower density. These increased the difficulty in microplastics removal. But biofilter technology still demonstrated the highest removal performance of microplastics (Fig. 2). Biofilter technology integrated physical and biological purification processes (Fig. 6C), and biofilm filtration and adsorption were the main mechanisms for microplastics removal. The microbe film growing on the surface of the inert filter material was in contact with the microplastics and increased the contact area between microplastics and microorgan isms. Excess microbes and retained microplastics were easy to be removed by backwashing in the ascendant water flow (Rocher et al., 2012). Microplastics are regarded as carriers of the microbes, so the occurrence of microplastics will influence the community and activity of microbial. Li et al. (2020) found that with the addition of PVC, the mi crobial community composition was reduced immediately and the number of operational taxonomic units decreased from 1665 to 1533. Subsequently, the number of operational taxonomic units increased to 1735. The percentage of each bacterial in the bacterial community slightly changed with the operation time. Therefore, the existence of microplastics PVC did not pose obvious operational taxonomic unit reduction and has an insignificant effect on the microbial community structure change. At the same time, it is gratifying that virgin micro plastics do not significantly affect the activities of ammonia oxidizing bacteria, nitrite oxidizing bacteria, and phosphorus accumulating or ganisms (Liu et al., 2019b). Therefore, the effect of microplastics on the performance of the bioreactor system should not be overemphasized. However, the toxicity of additives contained in the microplastic to bacteria was unclear. Subsequent research should consider the impact of microbe containing microplastics on conventional pollutant removal. Among them, the influence of microplastics on the function of microbe after adsorption of conventional pollutants also needs in-depth study. 4.5.2. Bioreactor system Fig. 6 showed the schematics of the bioreactor system in micro plastics removal. As shown in Fig. 6A, the bioreactor system removed microplastics mainly through the ingestion of microbe and the forma tion of sludge aggregates. In particular, domesticated activated sludges were likely to promote the accumulation of microplastics in WWTPs. Sludge containing microplastics was removed during the subsequent secondary settling process (Jeong et al., 2016). A2O is the most widely used bioreactor system in WWTPs (Fig. 2). However, it had a relatively poor microplastics removal efficiency owing to the sludge return. Some of the microplastics (20%) transferred into the sludge would flow back to the aqueous phase. Furthermore, the degradation of microplastics in A2O was quite slow. Previous studies reported several functional bacteria with microplastic degradation. The bacterial strain Rhodococcus degraded 6.4% of the PP polymer mass in 40 days (Auta et al., 2018). And Ideonella sakaiensis complete degraded PET film microplastics in 6 weeks (Yoshida et al., 2016). However, the existing hydraulic retention time (7–14 h) cannot achieve the effective degradation of microplastics only by microorganisms in WWTPs. Therefore, the conventional activated sludge method is not ideal for removing microplastics in WWTPs. Membrane bioreactor (MBR) technology has recently become a popular treatment process in WWTPs. It presents an excellent perfor mance in microplastics removal (removal efficiency 99.9%) (Talvitie et al., 2017a) due to a high mixed-liqueur suspended solids concentra tion (range from 6000 mg L− 1 to 10000 mg L− 1) (Dvořák et al., 2013). As shown in Fig. 6B, MBR technology integrated membrane separation and traditional activated sludge method. Most the microplastics were retained in the biofilm carrier side of the MBR system. This indicated that the adsorption effect was one of the major contributors to 4.5.3. Advanced oxidation Chlorination and UV-oxidation were the most popular advanced oxidation processes in WWTPs. Chlorine is a widely used as disinfection agent in WWTPs. Microplastics were not completely resistant to the attack of chlorine (Kelkar et al., 2019). The chlorination process increased the microplastic abundance because of the cracking of microplastics (Lv et al., 2019; Ruan et al., 2019). The schematic of microplastics degradation in chlorination was shown in Fig. 7A. Chlo rine potentially broke the existing bonds and introduced new bonds. The new chemical structure of high-density polyethylene (HDPE) in chlori nation disinfection was C-C-C asymmetrical chain, C-C-C symmetrical chain, CH2 twist, and CH2 bend (Kelkar et al., 2019), identified that intense chlorination might result in a force of compression on the Raman peaks. (Eichhorn et al., 2001). A new chlorine carbon bond (Cl-CH2-C-H) was formed in chlorination. Carbon-chlorine bonds might increase toxicity and hydrophobicity, which resulted in microplastics more easily adsorbing and accumulating harmful contaminants (Wang et al., 2018). Additionally, chlorine occurred on both the aliphatic and aromatic during PS degradation (Zebger et al., 2003). Raman shifting of the aliphatic C-H backbone towards a higher wavenumber (from 2901 cm− 1 to 2940 cm− 1) was observed. This shift signified the force of compres sion on the backbone bond towards higher energy (Eichhorn et al., 2001). Chlorination also changed the physical and chemical properties of microplastics due to its strong oxidizing nature (El-Shahawi et al., 2010). However, PP was resistant to chlorination. Even at high dosage and long exposure time, the change of chemical bond was scarcely observed (Kelkar et al., 2019). Due to competitive reactions and chlorine 9 W. Liu et al. Environment International 146 (2021) 106277 Fig. 7. The schematics of advanced oxidation technologies in microplastics removal. (A) Chlorination (Kelkar et al., 2019); (B) SEM images of the three types of virgin and UV-oxidation microplastics (Cai et al., 2018); (C) UV-oxidation; (D) PVC UV-oxidation reaction (Shi and Zhang, 2006). quenching, the coexistence of other pollutants, microorganisms and biofilms might alter the impact of microplastics structure in chlorination. The schematic of removal microplastics in UV-oxidation was shown in Fig. 7C. The UV-oxidation occurred on the surface of the micro plastics, resulting in the change of their topography and chemical characteristics (Cooper and Corcoran, 2010). Virgin microplastics pre sent relatively homogeneous and compact textures. After UV-oxidation, the surface of microplastics became relatively rough (Fig. 7B). Granular oxidation/flakes, cracks/flakes, and flakes were common situations of degradation for the PE, PP, and PS. However, microplastics with cracks and flakes were easy to break, which produced smaller and even nanoscale plastics (Cai et al., 2018). Cracks as an extension of fractures were considered as stress concentrators and fracture loci. The formation of flakes embrittlement occurs on the brittle surface areas or layer of the microplastics (Cooper, 2012). As shown in Fig. 7D, part of the peroxy free radicals was formed by the cleavage of C-C bonds and C-H bonds under the condition of UV irradiation (Cai et al., 2018; Gewert et al., 2015; Wang et al., 2020a). What’s more, UV irradiation could initiate hydroxyl groups (OH) and chromophore groups (including several – C), carbonyl groups (C– – O) and hydroperox unreacted monomers (C– ide groups (ROOH)) in the microplastic surfaces to form oxygencontaining free radicals and initiate the chain reactions. (Fig. 7D) (Cai et al., 2018; Singh and Sharma, 2008; Wang et al., 2020a; Zbyszewski and Corcoran, 2011). These peroxy free radicals produced by the UV irradiation would undergo secondary reactions to form crosslinking compounds (Fig. 7D). And the molecular chain with a carbonyl group will be broken to reduce the relative molecular mass (Cai et al., 2018). However, the intermediates and the toxicity of UV-oxidation micro plastics were less known. The impact of UV irradiation time and envi ronmental differences on microplastic degradation requires in-depth research. In addition, the influence of the salinity and dissolved organic matter on the degradation of microplastics in WWTPs also needs to be considered. 4.5.4. Membrane filtration Membranes with a uniform pore distribution have been widely applied during wastewater treatment. The schematic of removal microplastics in membrane filtration was shown in Fig. 8. Membrane filtration technologies intercepted microplastics in aqueous phase using different forms of membrane filtration (Baker, 2012). Microplastic particle size was larger than the ultrafiltration membrane pore size (nano-scale). Thus, microplastics were completely rejected by the ul trafiltration membranes (Ma et al., 2019b). Microplastics filtration led to a final water flux decline of 38% (Enfrin et al., 2020). This result showed the existence of interactions between microplastics and the membrane pores and surface. Microplastics were adsorbed within and onto the pores, or onto the membrane surface at a high rate. With the increase of exposure time, more and more microplastics were permeating across the membrane. For some special membranes, such as polysulfone mem branes, were relatively hydrophilic. Microplastics and polysulfone membranes were negatively charged and hydrophobic. Thus, attractive polar forces were counterbalanced by the repulsive electrostatic forces induced by the membrane surface charge and microplastics (Enfrin 10 W. Liu et al. Environment International 146 (2021) 106277 Fig. 8. The schematic of ultrafiltration technology in microplastics removal (Enfrin et al., 2020). et al., 2020). However, in order to ensure the long-term stable operation of the microfiltration and ultrafiltration membranes, fouling control must be strictly enforced (Kumar and Ismail, 2015). At the same time, intermo lecular repulsion of microplastics and the electrostatic interactions be tween microplastics and the membrane surface were the main mechanisms in microplastics removal by ultrafiltration technology. The electrostatic interactions between microplastics and ultrafiltration membranes were detrimental for the performance infiltration (Enfrin et al., 2020). Meanwhile, membranes suffered from surface fouling due to the formation of a concentration polarization layer during water transfer across the membrane (Enfrin et al., 2020). This concentration polarization induced the formation of the cake. It decreased the per formance of membrane filtration by adsorption and stacking of micro plastics or solutes onto the surface of the membrane (Baker, 2012). Therefore, efficient and stable cleaning procedures need further research to limit the influence of microplastics on membranes. capacity of WWTPs. These microplastics likely cause harm to aquatic organisms (Ma et al., 2020). Developing countries and areas with inef fective wastewater treatment processes should pay more attention to the microplastics pollution of aquatic environments. The microplastics in sludge are eventually retained in the soil envi ronment. Sludge is considered one of the most important sources of microplastics in the soil environment (Bläsing and Amelung, 2018). 43,000–63,000 and 30,000–44,000 tons of microplastics yearly entered European and North American agricultural soils, respectively (Nizzetto et al., 2016). The decomposition of these microplastics lasts up to 1,000 years (Tudor et al., 2019). Microplastics absorb toxic compounds and aggravate soil pollution (Li et al., 2019). The accumulation and trans portation of microplastics not only harms the growth of plants but also affects the functions and microbial communities in soil (Guo et al., 2020). The ecological toxicity effect and risk of compound pollution of microplastics and other pollutants necessitates further study. 6. Conclusions and future perspectives 5. Environmental toxicity and risk of microplastics A meta-analysis helps us to better understand the fate of micro plastics in WWTPs. The filter-based treatment process attained the highest microplastic removal efficiency. Fibers and microplastics with large particle sizes (0.5–5 mm) were easily separated by primary settling. PE and small-particle size microplastics (<0.5 mm) were easily trapped in the activated sludge and by bacteria in the WWTPs. The in teractions and removal mechanisms between microplastics and critical treatment technologies were quite different. Conventional flocculation interacted with microplastics via van der Waals forces, hydrogen bonds, or electrostatic forces in flocculation technology. The bioreactor system removed microplastics mainly through the ingestion of microbe and the formation of sludge aggregates. Advanced oxidation process affected the physical/chemical properties of microplastics, broke the existing bonds, and introduced new bonds. In membrane filtration technology, in teractions between microplastics and the membrane pores and surface made the microplastics easily adsorbed onto the membrane surface. Some of the microplastics removed from the above technologies were finally transferred into the sludge, the others released from the WWTPs posed environmental toxicity and risks. In current studies of the microplastics in WWTPs, certain problems need to be resolved in future research. To better evaluate the fate of the microplastics in WWTPs or other environmental media, further research should focus on the development of standardized sampling and analysis methods. Simultaneously, further research should prioritize the study of specific microplastics, especially in industrial zones. The influencing Microplastics enter aquatic and soil environments through waste water and sludge discharge. They are emerging pollutants as well as carriers of heavy metals and organic contaminants. Microplastics adsorbed heavy metals and polycyclic aromatic hydrocarbons could be ingested by benthic animals, leading to bioaccumulation in marine food chains (Foshtomi et al., 2019). Microplastics disrupt the soil structure and microbe metabolism and thus affect the water holding capacity of the soil (Machado et al., 2018). Certain properties (such as the structure) of land plants might improve the uptake of microplastics and other pollutants (He et al., 2018). The microplastics in the effluent from WWTPs ultimately converge in the aquatic environment (rivers and oceans). Primary microplastics are broken into secondary microplastics via physical, chemical, and bio logical treatment processes (González-Pleiter et al., 2019). As a conse quence, WWTPs are considered to be the main sources of secondary microplastics in the aquatic and soil environments. According to one investigation from a WWTP in China, the mass of microplastics dis charged into the water is less than <10 kg per day (Section 4.3), but due to its low density and small volume, the number of microplastics frag ments is still high (Lv et al., 2019). As listed in Tables 1 and S1, the abundance of the microplastics in the effluent is 5.00 × 105–1.39 × 1010 particles on a daily basis (mean value: 7.20 × 109 particles; median value: 2.00 × 108 particles). In other words, billions of microplastic particles are discharged into rivers every day due to the high daily 11 Environment International 146 (2021) 106277 W. Liu et al. factors of the treatment processes in removing microplastics in the WWTPs also requires in-depth research, such as hydraulic retention time, salinity and dissolved organic matter. In addition, the impact of conventional pollutant removal, reaction intermediates and their toxicity generated by the existing treatment process in the removal of microplastics were less known. Particularly, the toxicity of additives contained in the microplastic to bacteria. Microplastic-targeted treat ment technologies are also urgently needed to avoid emissions into aquatic and soil environments. Moreover, the potential impact of the subsequent utilization of sludge on the soil environment should be investigated in the future. In particular, research on the effects of different polymers on plant roots is necessary. 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