Mohsen Padervand
email: padervand@maragheh.ac.ir

Eric Lichtfouse
email: eric.lichtfouse@gmail.com

Didier Robert
email: didier.robert@univ-lorraine.fr

Chuanyi Wang
email: wangchuanyi@sust.edu.cn

Removal of microplastics from the environment. A review

Keywords: Microplastic pollution, Environment, Removal method

The production of fossil fuel-derived, synthetic plastics is continually increasing, while poor plastic waste management has recently induced severe pollution issues. Microplastics are plastic particles smaller than 5 mm. Microplastics are ubiquitous and slowly-degrading contaminants in waters and soils. Microplastics have long residence time, high stability, high potential of being fragmented and can adsorb other contaminants. Many aquatic species contain microplastics, which are in particular easily accumulated by planktonic and invertebrate organisms. Then, microplastics are transferred along food chains, leading to physical damages, decrease in nutritional diet value and exposure of the living organism to pathogens. Raw plastics contain chemical additives such as phthalates, bisphenol A and polybrominated diphenyl ethers that may induce toxic effects after ingestion by living organisms. Furthermore, the adsorption capability of microplastics makes them prone to carry several contaminants. Methods to remove microplastics from water and other media are actually needed. Here, we review microplastics occurrence, transport, raw polymers and additives, toxicity and methods of removal. Removal methods include physical sorption and filtration, biological removal and ingestion, and chemical treatments. Mechanisms, efficiency, advantages, and drawbacks of various removal methods are discussed.

Introduction

The global production of plastics has highly increased since 1950 to improve human life and reached almost 381 million tons in 2015 [START_REF] Ritchie | Plastic pollution[END_REF]. This increase has, however, induced global plastic pollution, making plastics pollutants of concern [START_REF] Macarthur | Microplastics baseline surveys at the water surface and in sediments of the North-East Atlantic[END_REF]. Microplastics are plastics with size lower than 5 mm, originating from the exfoliation and degradation of many types of plasticbased products released into ecosystems [START_REF] Zhang | Microplastic pollution in China's inland water systems: a review of findings, methods, characteristics, effects, and management[END_REF]. Microplastics has been reported in ocean sediments [START_REF] Van Cauwenberghe | Microplastic pollution in deep-sea sediments[END_REF], urban and rural areas [START_REF] Hirai | Organic micropollutants in marine plastics debris from the open ocean and remote and urban beaches[END_REF], freshwaters [START_REF] Faure | Plastic pollution in Swiss surface waters: nature and concentrations, interaction with pollutants[END_REF] and seawaters [START_REF] Law | Distribution of surface plastic debris in the eastern Pacific Ocean from an 11-year data set[END_REF]. Most reports suggest an accumulation of microplastics in aquatic environments, and, as a consequence, a higher exposure of living organisms to microplastics and their degradation by-products [START_REF] Andrady | Microplastics in the marine environment[END_REF][START_REF] Sun | Microplastics in wastewater treatment plants: detection, occurrence and removal[END_REF].

Microplastics are categorized as primary microplastics, which are raw materials used in domestic and personal care products, and secondary microplastics arising from the degradation of raw plastic particles by physical, chemical, and biological processes in the environment [START_REF] Galgani | Marine litter within the European marine strategy framework directive[END_REF]). Long-term durability due to their polymeric structure and easy transport between different habitats make microplastics of high concern for biologists and environmentalists (Fig. 1). Major raw polymers include polyethylene terephthalate (PET), polyurethane (PU), polystyrene (PS), polyvinylchloride (PVC), polypropylene (PP), polyesters, polyethylene (PE) and polyamide (PA, nylon). Poor plastic waste management has resulted in ubiquitous microplastics occurrence [START_REF] Gilani | Tautomerism, solvatochromism, preferential solvation, and density functional study of some heteroarylazo dyes[END_REF][START_REF] Thompson | Microplastics in the marine environment: sources, consequences and solutions[END_REF]. Several reports show that long-term exposure to microplastics causes chronic toxicity, yet there is no evidence on their acute fatal effects (Li et al. 2018a;[START_REF] Sussarellu | Oyster reproduction is affected by exposure to polystyrene microplastics[END_REF]. Microplastic toxicity is controlled by different routes depending upon their chemical structure and additives used as linkage during polymerization [START_REF] Meeker | Phthalates and other additives in plastics: human exposure and associated health outcomes[END_REF][START_REF] Sussarellu | Oyster reproduction is affected by exposure to polystyrene microplastics[END_REF]. As an example, polystyrene microplastics are able to be transferred in blood, causing reproductive disruption in marine filter feeders [START_REF] Law | Distribution of surface plastic debris in the eastern Pacific Ocean from an 11-year data set[END_REF].

To our best knowledge, this is the first review on microplastics removal. We discuss microplastic additives, occurrence, transport and toxicity, then we review removal methods. Removal methods include sorption and filtration, removal based on chemical phenomena, and biological ingestion treatments. Advantages, disadvantages and efficiency of different methods are compared at the end.

Microplastic sources, transport, polymers and additives

Microplastic sources and occurrence

Microplastics can be found worldwide in coastal regions and aquatic ecosystems in various size fractions due to the transport phenomena including wind and ocean currents. Primary sources are household sewage discharge including polymeric particles from cosmetic and cleaning products, feedstocks used to manufacture plastic products, and plastic pellets or powders used for air blasting [START_REF] Jiang | Occurrence of microplastics and its pollution in the environment: a review[END_REF]. Progressive fragmentation of larger plastic items under the atmospheric conditions, e.g., by mechanical degradation and UV light exposure, thus contributing to the entrance of considerable amounts of microplastics to the environment, is the secondary source of microplastics [START_REF] Eriksen | Plastic pollution in the world's oceans: more than 5 trillion plastic pieces weighing over 250,000 tons afloat at sea[END_REF]. This increases plastic debris availability for being ingested by a large variety of organisms and highlights the appearance of further environmental hazards [START_REF] Thompson | Plastics, the environment and human health: current consensus and future trends[END_REF].

Wastewater treatment plants are also a major source of microplastics release [START_REF] Browne | Accumulation of microplastic on shorelines worldwide: sources and sinks[END_REF][START_REF] Long | Microplastic abundance, characteristics, and removal in wastewater treatment plants in a coastal city of China[END_REF]. Whereas large plastic particles are efficiently removed during wastewater treatment, microplastics often bypass the treatment units, thus entering and accumulating in the aquatic environment [START_REF] Murphy | Wastewater treatment works (WwTW) as a source of microplastics in the aquatic environment[END_REF]. Noteworthy, a large number of water treatment plants are located near the ocean and seawater, thus inducing a high microplastic release source. For instance, in mainland China, about 1873 wastewater plants (56%), out of 3340, with 78 × 10 6 m 3 /day of treatment capacity, are located in coastal regions where their effluents can be directly or indirectly discharged into aquatic ecosystems [START_REF] Jin | Current state of sewage treatment in China[END_REF]). To address this issue, many researchers are investigating the fate, occurrence, detection and removal of microplastics in the water treatment plants [START_REF] Beljanski | Efficiency and effectiveness of a low-cost, selfcleaning microplastic filtering system for wastewater treatment plants[END_REF][START_REF] Carr | Transport and fate of microplastic particles in wastewater treatment plants[END_REF][START_REF] Sun | Microplastics in wastewater treatment plants: detection, occurrence and removal[END_REF].

Microplastic transport

Sea and ocean are viewed as the major sinks for microplastics, whereas freshwaters and terrestrial environments are the main sources. Indeed, early research found that microplastic litter reaching oceans by rivers contains particles found in soils (Horton et al. 2017a). This implies that freshwaters and soils are also sinks of microplastics, as evidenced by high concentrations of microplastics in some terrestrial and freshwater areas [START_REF] Nizzetto | Are agricultural soils dumps for microplastics of urban origin?[END_REF]. The long-term durability of microplastic fibers found in deeper layers (~ 25 cm) of agricultural soils treated by sewage sludge as fertilizer [START_REF] Zubris | Synthetic fibers as an indicator of land application of sludge[END_REF], suggests a gradual transport in solid media, then further accumulation at depth; thus making agricultural and forest soils more likely to retain microplastics compared to urban areas [START_REF] Lwanga | Incorporation of microplastics from litter into burrows of Lumbricus terrestris[END_REF].

As rivers carry a huge volume of plastic particles over the large distances, microplastics probably settle out along with sinking sediments, particularly where flow energy drops, for instance in retard-moving riverbeds. Accordingly, futher sediment deposition of microplastics in lakes where water flow is the lowest and sedimentation rate is high, should induce high accumulation [START_REF] Corcoran | Hidden plastics of Lake Ontario, Canada and their potential preservation in the sediment record[END_REF].

The shape diversity, small size, lightweight and low density of microplastics contribute to their widespread transport and facile dispersal across large distances on land and within aquatic systems by storm sewers, wind and other natural currents [START_REF] Horton | Microplastics: an introduction to environmental transport processes[END_REF]. Larger size and higher density result in facile sinking and sediment deposition of the microplastics (Horton et al. 2017b). Furthermore, irregularly shaped microplastics with jagged geometry and sharp ends are more likely retained underwater, rather than returning to the surface, whereas spherical particles show a higher tendency to stay at the surface [START_REF] Ballent | Physical transport properties of marine microplastic pollution[END_REF][START_REF] Lagarde | Microplastic interactions with freshwater microalgae: heteroaggregation and changes in plastic density appear strongly dependent on polymer type[END_REF].

Microplastics transport pathways in the air are not fully understood [START_REF] Horton | Microplastics: an introduction to environmental transport processes[END_REF]. Noteworthy, in the air, there are few dispersal boundaries, compared to water systems. Nonetheless, microplastics transport within the atmosphere is not totally independent of aquatic and terrestrial pollutions, and here further investigations are needed to elucidate the mechanisms [START_REF] Dris | Synthetic fibers in atmospheric fallout: a source of microplastics in the environment?[END_REF].

As another major concern, due to their hydrophobicity and high surface area/volume ratio, microplastics are highly susceptible to sorb and carry persistent organic pollutants such as polychlorinated biphenyls (PCB), dichlorodiphenyltrichloroethane (DDT) and polyaromatic polycyclic aromatic hydrocarbons (PAH), which can be subsequently transferred to coastal regions and be desorbed inside living organisms [START_REF] Browne | Microplastic moves pollutants and additives to worms, reducing functions linked to health and biodiversity[END_REF]. Consequently, the concentration of organic pollutants in coastal areas is expected to increase several orders of magnitude as a result of pollutant transport by microplastics. Microplastic morphology and transport are thus major characteristics controlling the other waterborne pollutants [START_REF] Cole | Microplastics as contaminants in the marine environment: a review[END_REF].

Microplastic raw polymers and additives

Polymeric ingredients of primary microplastics mainly include polyethylene, polypropylene and polystyrene, depending upon the type of the products manufactured by the factory; while secondary microplastics are predominantly made of polyester, acrylic and polyamide, forming fibers in the environment [START_REF] Jiang | Occurrence of microplastics and its pollution in the environment: a review[END_REF]. The microplastic number in the inland freshwaters of Wuhan in China ranged between 1660.0 ± 639.1 and 8925 ± 1591 numbers/m 3 ; here the major types were polyethylene terephthalate and polypropylene [START_REF] Wang | Microplastics pollution in inland freshwaters of China: a case study in urban surface waters of Wuhan, China[END_REF]. Low-density polyethylene has been also identified as the dominant type of microplastics.

Microplastics contain a large variety of chemical additives such as bisphenol A, phthalates and polybrominated diphenyl ethers, which are used during raw plastic synthesis to improve plasticity [START_REF] Besseling | Nanoplastic affects growth of S. obliquus and reproduction of D. magna[END_REF][START_REF] Murphy | Additives for plastics handbook[END_REF]). These additives are endocrine disruptors, and thus may exhibit toxic effects upon release. The concentration of such plasticizers in plastic debris of remote and urban beaches is up to 35 ng/g in remote beaches and up to 700 ng/g in urban beaches for bisphenol A; between 0.1 and 400 ng/g in remote beaches and up to 9900 ng/g in urban beaches for polybrominated diphenyl ethers; and up to 3940 ng/g for phthalates [START_REF] Hirai | Organic micropollutants in marine plastics debris from the open ocean and remote and urban beaches[END_REF]. These plastic additives have been detected in most microplastic polymers [START_REF] Jiang | Occurrence of microplastics and its pollution in the environment: a review[END_REF]. Researchers also reported the leaching of bisphenol A and nonylphenol from silicone and polycarbonate microplastics [START_REF] Fasano | Migration of phthalates, alkylphenols, bisphenol A and di (2-ethylhexyl) adipate from food packaging[END_REF]. Accumulation of such chemicals in human bodies through biological phenomena is also reported [START_REF] Talsness | Components of plastic: experimental studies in animals and relevance for human health[END_REF]). The most alarming exposure route to microplastics for human is food, where the adverse effects of the chemical additives and mechanism of entrance to the organs are still unexplored [START_REF] Wright | Plastic and human health: a micro issue?[END_REF]. Accordingly, many efforts must be devoted to finding efficient strategies to abate the presence of microplastics in the environment. While there have been published reports on characterizing sources, occurrence, fate, methods for detection, and environmental effects; to date, few research and review papers have discussed removal processes of microplastics from contaminated systems.

Toxicity of microplastics

Toxicity from the chemical structure

The potential toxicity of microplastics arises from unreacted monomers, oligomers and chemical additives leaked from the plastic in the long rub [START_REF] Thompson | Lost at sea: where is all the plastic?[END_REF]). Monomers and oligomers are both able to migrate from food packaging materials [START_REF] Piringer | Plastic packaging materials for food: barrier function, mass transport, quality assurance, and legislation[END_REF]. As the concentration of the residuals reaches specific limits, they can be potentially absorbed by human bodies via different pathways. For instance, the presence of polystyrene residuals in food materials is reported to cause serious health issues, while epoxy resins made of bisphenol A are absorbed by living tissues, then interfer with the rate of cell division [START_REF] Lau | Contamination in food from packaging material[END_REF].

Chemical additives are used during polymers manufacturing for improving the products performance. Additives include functional additives such as plasticizers, heat stabilizers, flame retardants, antioxidants, colorants, e.g. soluble azo-colorants and pigments, fillers such as kaolin and clay, and reinforcements, e.g. carbon and glass fibers. These additives are another source of toxicity. For example, researchers found that the release level of some phthalates from baby bottles was in the range of 50-150 μg/kg of food content after the contact time of 120 min at 70 °C [START_REF] Simoneau | Identification and quantification of the migration of chemicals from plastic baby bottles used as substitutes for polycarbonate[END_REF]). The release level of bisphenol A from food packaging items was estimated to be in the range of 100-800 ng/L, while the values were in the range of μg/L for some phthalates under the same conditions [START_REF] Fasano | Migration of phthalates, alkylphenols, bisphenol A and di (2-ethylhexyl) adipate from food packaging[END_REF]. Most of these additives are not chemically bound to the bulk plastic structures, implying easier release. [START_REF] Nobre | Assessment of microplastic toxicity to embryonic development of the sea urchin Lytechinus variegatus (Echinodermata: echinoidea)[END_REF] studied the toxicity of raw and beachstranded microplastics on the development of embryos of Lytechinus variegatus, simulating leaching of the chemical additives into the water column and interstitial water by assays of elutriate and pellet-water interface, respectively. They found that raw microplastics induced more toxicity, enhancing anomalous embryonic development by 58.1% and 66.5% for the former and latter evaluation method, respectively. Their results also implied that the leaching of chemical compounds strongly depends upon the media compartment in which microplastics accumulate, and upon the exposure pathway. [START_REF] Hahladakis | An overview of chemical additives present in plastics: migration, release, fate and environmental impact during their use, disposal and recycling[END_REF] reviewed migration and release rate, fate, and potential toxicity effects of additives on organisms and environment. The release of volatile compounds, e.g., benzene, toluene, ethylbenzene, styrene and methylene chloride, from plastics can also contribute to chronic health effects [START_REF] Andrady | The plastic in microplastics: a review[END_REF][START_REF] Huff | Clarifying carcinogenicity of ethylbenzene[END_REF][START_REF] Wexler | Encyclopedia of toxicology[END_REF].

Toxicity from physical properties

Microplastics exert damage through the effect of a relatively large surface area/volume ratio. They absorb hydrophobic pollution from water, then carry this pollution to other habitats [START_REF] Setälä | Ingestion and transfer of microplastics in the planktonic food web[END_REF]. A study of the effect of phenanthrene-loaded low-density polyethylene glycol microplastics on biomarker responses in juvenile African catfish revealed significant tissue changes in the liver and brain of the organism [START_REF] Karami | Virgin microplastics cause toxicity and modulate the impacts of phenanthrene on biomarker responses in African catfish (Clarias gariepinus)[END_REF].

The ingestion of microplastics by biota is a common way to induce toxicological effects [START_REF] Hämer | Fate of microplastics in the marine isopod Idotea emarginata[END_REF]. Polystyrene microplastics enhance the bioavailability of fluoranthene compounds to marine mussels (Mytilus spp.) after 7 days of exposure under controlled experimental conditions [START_REF] Paul-Pont | Exposure of marine mussels Mytilus spp. to polystyrene microplastics: toxicity and influence on fluoranthene bioaccumulation[END_REF]. These results mean that ther is a higher fluoranthene concentration in mussels exposed to fluoranthene-loaded microplastics than those exposed to pure fluoranthene. Highest levels of antioxidant markers and histopathological damages were also observed for the former case. They explained the mechanism by interactions between the cell wall components of the marine mussels, e.g. p-glycoprotein, involved in pollutant excretion, and the microplastics surface. [START_REF] Zhang | Toxic effects of microplastic on marine microalgae Skeletonema costatum: interactions between microplastic and algae[END_REF] investigated the adverse effects of microplastics on the photosynthesis of the marine microalgae Skeletonema costatum. They found that the maximum growth inhibition ratio reached up to 39.7% after 96 h of exposure to microplastics with average diameter of 1 μm. Their results show a drastic decrease in chlorophyll content (20%) and photosynthetic efficiency (32%) after exposure to high concentration of microplastic (50 mg/L), leading to negative effects on microalgae growth. According to the results of morphological studies and scanning electron microscopy (SEM), they proposed both adsorption and aggregation of microplastics on the outer surface of microalgae as the most probable mechanism of toxicity.

Size dependency of microplastics toxicity was also confirmed by [START_REF] Lu | Uptake and accumulation of polystyrene microplastics in zebrafish (Danio rerio) and toxic effects in liver[END_REF] who investigated the exposure effects of polystyrene microplastics to zebra fish. They stated that a 7-day exposure resulted in accumulating the microplastics with size of 5 μm in liver, gill and gut, while those with size of 20 μm were just found in fish gill and gut. Moreover, lipid accumulation and inflammation of liver, oxidative stress, and adverse alterations in the metabolism profile of the fish liver were the main toxicity outcomes.

The shape and texture of the ingested microplastics also influence their toxicity and absorption capability. According to [START_REF] Au | Responses of Hyalella azteca to acute and chronic microplastic exposures[END_REF], polypropylene microplastic fibers were more toxic than polyethylene microplastic spherical particles to the freshwater amphipod, Hyalella azteca. They attributed this to the longer residence time of the fibers in gut, which modifies the ability of food processing, thus leading to serious changes in sublethal endpoints.

Toxicity from microorganisms carried by microplastics

The potential of microplastics to carry pathogenic bacteria has been explored by [START_REF] Kirstein | Dangerous hitchhikers? Evidence for potentially pathogenic Vibrio spp. on microplastic particles[END_REF]. They observed Vibrio parahaemolyticus bacterial strains on some polyethylene, polypropylene and polystyrene marine microplastic particles gathered from North Sea. They highlighted the need for further consideration of health impacts of microbial assemblages in microplastics. A 10-day exposure to five types of ~ 70 μm microplastics led to intestinal damage including splitting of enterocytes and cracking of villi in zebrafish Danio rerio and nematode C. elegans, as model organisms of freshwater [START_REF] Lei | Microplastic particles cause intestinal damage and other adverse effects in zebrafish Danio rerio and nematode Caenorhabditis elegans[END_REF]. They also demonstrated that 2-day exposure of 5.0 mg/m 2 of microplastics considerably reduced calcium levels and survival rates, and inhibited body length and reproduction of C. elegans. They suggested that oxidative stress and intestinal damages are the main toxicity effects of microplastics. [START_REF] Prata | Airborne microplastics: consequences to human health?[END_REF] reviewed the potential toxicity of airborne microplastics and adverse effects of their low environmental concentrations on human health. They discussed the diseases aroused from airborne microplastics and pathophysiological mechanisms of toxicity including dust overload, oxidative stress, translocation, and gene mutation. They proposed that exposure to low atmospheric concentrations can contribute to incidence of cardiovascular, respiratory, and interstitial lung diseases.

Overall, the toxicity of microplastics arises from raw chemical additives, adsorption and transport of pollutants and microbes, and release to life and the environment. Toxicity depends physically on size and shape of microplastics. Biota and humans are affected by toxic effects of microplastics via mechanisms including sorption and aggregation in different organs, ingestion, and exertion of physical damages and disturbing the life systems. Research is needed to clarify how microplastics induce tissue changes and pathological disorders.

Removal of microplastics using sorption and filtration methods

Adsorption on green algae

The presence of microplastics in aquatic environments is expected to be more critical than other pollutants due to several harmful effects and death of organisms, e.g. fishes, mammals, marine birds and reptiles, arising from their entanglement and bioaccumulation [START_REF] Cole | Microplastics as contaminants in the marine environment: a review[END_REF][START_REF] Graham | Deposit-and suspension-feeding sea cucumbers (Echinodermata) ingest plastic fragments[END_REF][START_REF] Gregory | Environmental implications of plastic debris in marine settings-entanglement, ingestion, smothering, hangers-on, hitch-hiking and alien invasions[END_REF]. Their persistence and low degradability call for removal methods. Microplastics are generally categorized as persistent materials but they degrade more of less depending upon their nature and chemical structure. Microplastics with half-life times lower than the values defined in terms of REACH criteria for persistency (Table 1) can be considered as degradable, and do not pose a threat to the environment [START_REF] Verschoor | Towards a definition of microplastics: Considerations for the specification of physico-chemical properties[END_REF]. Microplastics are prone to sorb a large variety of waterborne contaminants on the surface, carry them and desorb into the new habitats [START_REF] Rios | Persistent organic pollutants carried by synthetic polymers in the ocean environment[END_REF]. Their large surface area/ volume ratio makes adsorption of other contaminants likely. [START_REF] Sundbaek | Sorption of fluorescent polystyrene microplastic particles to edible seaweed Fucus vesiculosus[END_REF] studied the adherence behavior of fluorescent microplastic particles on the surface of an edible marine microalgae, seaweed, named Fucus vesiculosus. The diameter size of the polystyrene microplastics was ~ 20 μm, while the plant cells of the sorbent contained very narrow microchannels to restrict the translocation of polystyrene microplastics into the tissues. The results revealed a high sorption of microplastics (~ 94.5%), mainly near the cut surfaces of the seaweed, which is explained by the role of released alginate compounds from cell walls in the cut regions. Indeed, because of the gelatinous characteristics of this anionic polysaccharide substance, alginate is able to improve the adherence of polystyrene particles on the seaweed's surface [START_REF] Martins | Sex-biased gene expression in the brown alga Fucus vesiculosus[END_REF]. This paper and the other researches on microalgae capabilities to sorb tiny plastic particles accentuate the effective role of microplastics' surface charge and microalgae's surface characters [START_REF] Bhattacharya | Physical adsorption of charged plastic nanoparticles affects algal photosynthesis[END_REF][START_REF] Nolte | The toxicity of plastic nanoparticles to green algae as influenced by surface modification, medium hardness and cellular adsorption[END_REF]. Investigating the adsorption of 20-500 nm polystyrene particles onto unicellular green algae, Pseudokirchneriella subcapitata, [START_REF] Nolte | The toxicity of plastic nanoparticles to green algae as influenced by surface modification, medium hardness and cellular adsorption[END_REF] concluded that positively charged polystyrene microplastics are more efficiently adsorbed on the algae's surface than those with negative charge.

Overall, the sorption of microplastics on algae surface strongly depends on particles' surface charge. Positively charged microplastics have higher tendency to be sorbed more efficiently, which is explained by the presence of an anionic polysaccharide in the algal chemical structure. Li et al. (2018b) reported the use of dynamic membranes for the efficient removal of microplastics from a synthetic wastewater (Fig. 2). They investigated the effect of influent flux and particles concentration on the removal efficiency of dynamic membranes formed on a diatomite platform with 90 μm of supporting mesh during filtration of the synthetic wastewater. Excellent filtration of microplastices was obtained in 20 min by decreasing the turbidity from 195 NTU for the influent to less than 1 for the effluent [START_REF] Ersahin | Impact of anaerobic dynamic membrane bioreactor configuration on treatment and filterability performance[END_REF][START_REF] Horton | Microplastics: an introduction to environmental transport processes[END_REF]. Dynamic membrane formation is facilitated at higher influent fluxes and microplastics concentrations. [START_REF] Ward | Marine microplastic removal tool[END_REF] designed an efficient microplastic removal tool based on polymer coatings as an elongated mesh screen. He claimed that the tool has good durability and has the advantage of being easily fabricated from commodious materials. Other tool advantages included the absence of electrical power and mechanical devices. Membrane bioreactors, however, exhibit higher capacitie than simple dynamic membranes for the removal of microsized plastics [START_REF] Lares | Occurrence, identification and removal of microplastic particles and fibers in conventional activated sludge process and advanced MBR technology[END_REF]Talvitie et al. 2017a). [START_REF] Knoblock | Membrane biological reactor system for treatment of oily wastewaters[END_REF] studied the purification capability of a coupled system, taking advantage of porous membranes in combination with biological processes. Successful removal of a large variety of complex industrial wastewaters by membrane bioreactors confirms the suitability of this technology for the treatment for high-strength contaminants such as polymeric debris and microplastics [START_REF] Gurung | Incorporating submerged MBR in conventional activated sludge process for municipal wastewater treatment: a feasibility and performance assessment[END_REF]. Talvitie et al. (2017a) investigated the removal of various types of microplastics from wastewater treatment plant effluents using advanced final-stage treatment technologies including membrane bioreactor, disk filter, rapid sand filtration, and dissolved air floating (Fig. 3). They concluded that the membrane bioreactor eliminated 99.9%, from 6.9 to 0.005 microplactic particle per L (Table 2). They also showed that membrane bioreactor, rapid sand filtration and dissolved air floating removed microplastics of any size, even the smallest size fractions of 20-100 μm. Moreover, the removal efficiency did not depend upon the microplastics shape; particularly, textile fibers which were predominant in both influents and effluents during the treatment, were efficiently removed. Analysis of the samples by Fourier transform infrared spectroscopy (FTIR) indicated that the membrane bioreactor highly decreased the number of polymers in the final effluent, too, which highlights the good sorption capacity of the setup to trap microplastics of various chemical structures.

Removal using membrane technology

Membrane technologies were successfully used to remove microplastics from polluted aquatic environments. The removal efficiency over the membranes particularly depends on its durability, influent flux, size, and concentration of the microplastics. The combination of porous membranes with biological processes could enhance the removal efficiency up to 99.9%. [START_REF] Lares | Occurrence, identification and removal of microplastic particles and fibers in conventional activated sludge process and advanced MBR technology[END_REF] recently studied the performance of a municipal wastewater treatment plant operating based on a pilot-scale, combined membrane bioreactor-conventional activated sludge methodology for the removal of microplastics. Their study took 3 months and sampling campaigns were every 2 weeks. The wastewater samples were collected from a municipal water treatment plant located next to the city center of Mikkeli in Finland. Including an aeration tank, for mixing the wastewater with air to activate microorganisms, and a sedimentation tank, where the sludge is separated from the treated wastewater, for the biological degradation and secondary purification, respectively, a conventional activated sludge system is also expected to improve the activity [START_REF] Gurung | Incorporating submerged MBR in conventional activated sludge process for municipal wastewater treatment: a feasibility and performance assessment[END_REF]. Results revealed a better removal of microplastics (99.4%) using a membrane bioreactor compared to the conventional activated sludge treatment system (98.3%). Microplastics concentration in the final effluent of the former system was estimated to be 0.4 ± 0.1 MP/L, which was lower in comparison with that evaluated for the later process (1.0 ± 0.4 MP/L). They also stated that the main reason for observing a narrow range of final microplastics concentrations is probably due to opting for different processing steps and wastewater samples than previous works [START_REF] Leslie | Microplastics en route: field measurements in the Dutch river delta and Amsterdam canals, wastewater treatment plants, North Sea sediments and biota[END_REF][START_REF] Mintenig | Identification of microplastic in effluents of waste water treatment plants using focal plane array-based micro-Fourier-transform infrared imaging[END_REF][START_REF] Murphy | Wastewater treatment works (WwTW) as a source of microplastics in the aquatic environment[END_REF].

Removal using advanced filtration technologies in wastewater treatment plants

11 wastewater treatment plants in Changzhou, China, were studied for their efficiency to remove microplastics by following the abundance, color, shape and dimensional changes during the removal steps [START_REF] Ma | Removal characteristics of microplastics by Fe-based coagulants during drinking water treatment[END_REF]. All plants that used several treatment steps such as subsequent tanks for floating and sedimentation, and filtration processes, eliminated more than 90% of microplastics from the influents, with a final removal efficiency reaching 97.15%. The most important reasons that cause variation in removal efficiencies showing the decrease in turbidity with time when microplastics are removed (Li et al. 2018b). TMP transmembrane pressure, NTU nephelometric turbidity unit might be daily processing volume, different raw water and type of treatment processes. Large-size microplastics were less abundant in the effluents, in agreement with previous reports [START_REF] Horton | Microplastics: an introduction to environmental transport processes[END_REF]. In addition, fiber rayon and polyethylene terephthalate were the main ingredients of these microplastics, and reached the highest removals.

The microplastics removal methodology used by municipal sewage treatment plants was studied by researchers in Beijing, China [START_REF] Yang | Removal of microplastics in municipal sewage from China's largest water reclamation plant[END_REF]). The influents were initially treated by a series of processes including aerated grit chamber, primary and secondary sedimentation tanks following A 2 O treatments, e.g., anaerobic, anoxic and aerobic. Finally, advanced treatment processes including denitrification, ultrafiltration, ozonation and ultraviolet are applied to complete the cycle and remove microplastics from wastewater. The results of FTIR analysis revealed the presence of 18 polymers in the effluent, in which polyethylene terephthalate and polyester had the highest abundances of 42.26 and 19.1%, respectively. 58.84% of the microplastics population in influents was removed during the primary step of aerated grit treatment, while the removal efficiency reached to 71.67% following the advance treatment processes. Although the overall removal efficiency in the present sewage treatment plant, of 95.16%, was outstandingly less than that of for membrane bioreactors, of 99.9%, it was comparable to the efficiency of dissolved air flotation technologies and sand filters (Talvitie et al. 2017b). Noteworthy, the treatment processes currently employed at sewage treatment plants are not necessarily designed for the removal of microplastics. Regardless, these processes are able to eliminate a large portion of microplastics from the wastewater. The researchers finally estimated that sewage treatment plants release almost 0.59 × 10 9 items of microplastic into the aquatic ecosystems.

Overall, sorption and filtration methodologies show a good efficiency for the treatment for microplastics-containing wastewater, mainly in combination with other procedures such as biological and sedimentation processes. To achieve higher removal efficiency, membrane bioreactors are simultaneously used with other advanced physical-chemical treatments in most wastewater treatment plants all over the world, but it has still been confirmed that these plants act as microplastics sources to aquatic environment.

Chemical methods to treat microplastics

Many wastewater treatment plants worldwide use coagulation and agglomeration processes to form enlarged contaminant particles that are easier to separate [START_REF] Hu | Effect of aluminum speciation on arsenic removal during coagulation process[END_REF][START_REF] Lee | Development, characterization and the application of hybrid materials in coagulation/ flocculation of wastewater: a review[END_REF][START_REF] Shirasaki | Effect of aluminum hydrolyte species on human enterovirus removal from water during the coagulation process[END_REF]). These processes involve Fe-and Al-based salts and other coagulants to bind tiny particles via uptake-complexation mechanisms initiated by a ligand exchange, thus forming strong bonds between waste particles [START_REF] Chorghe | Boron removal from hydraulic fracturing wastewater by aluminum and iron coagulation: mechanisms and limitations[END_REF]). Ariza-Tarazona et al. ( 2019) recently studied the removal of polyethylene microplastics using iron and aluminum salt coagulants and ultrafiltration (Fig. 4). The experiments were carried out under different concentrations of Al 3+ and Fe 3+ ions, and the results indicated that Al 3+ has better performance than Fe 3+ . Also, the microplastic removal efficiency was scarcely modified by the pH of the solution at low concentration of Al coagulant source, 0.5 mM, whereas removal efficiency decreased by increasing the pH, particularly for small-sized microplastics, of diameter lower than 0.5 mm. They found that polyacrylamide (PAM), an enhancing coagulation agent, increased the removal efficiency for small microplastics much better than for large particles under high Al dosage of 5 mM. This highlights the greater growth rate of small microplastics in the presence of cationic polyacrylamide. The growth rate was highly enhanced when anionic polyacrylamide was used for the removal efficiency for the smaller microplastics (d < 0.5 mm): here the removal was raised from 25.83% without polyacrylamide to 61.19% with 15 mg/L polyacrylamide, while it just increased from 4.27% to 18.34% for larger microplastics, of 2-5 mm diameter.

Ariza-Tarazona et al. ( 2019) also explored the removal of the microplastics with the same method but using FeCl 3 •6H 2 O coagulation agent (Fig. 5). They found that the Fig. 4 Microplastic removal by coagulation, sedimentation and ultrafiltration (UF) [START_REF] Ma | Removal characteristics of microplastics by Fe-based coagulants during drinking water treatment[END_REF]. The effect of anionic polyacrylamide (PAM), pH and the formation of Al-based flocs on the removal efficiency is well represented removal at neutral pH enhanced as the coagulant concentration increased and, this trend being clearer for small microplastics, of d lower than 0.5 mm. Similarly, the removal efficiency was accelerated at high pH and high concentration, 2 mM, of the coagulant and for lower size of the microplastics. However, in this case and under high dosage of the coagulation agent, 2 mM, anionic polyacrylamide acted much better than the cationic one to improve the removal rate of polyethylene microplastics. This can be mechanistically explained based on facile formation of Fe-based flocs during the coagulation process, during which anionic polyacrylamide makes the products dense enough to be easily trapped and separated.

Researchers have also used the robust and environmentally compatible electrocoagulation technique [START_REF] Perren | Removal of microbeads from wastewater using electrocoagulation[END_REF], which allows sludge minimization, energy efficiency, cost-effectiveness, and flexibility to automation, to remove the polyethylene microplastics in a stirred-tank batch reactor (Fig. 6). In situ formation of the metal hydroxide coagulants is initiated by the reaction of metal ions such as Fe 2+ and Al 3+ released from sacrificial electrodes in a water stream with hydroxyl anions of the media; Eqs. 1-4 display the anodic and cathodic reactions, respectively:

(1)

M (s) → M n+ (aq) + ne - (2) 2H 2 O (l) → 4H + (aq) + O 2(g) + 4e - (3) M n+ (aq) + ne -→ M (s) (4) 2H 2 O (l) + 2e -→ H 2(g) + 2OH -
The produced coagulants break up the colloids and stabilize the suspended microparticles surface charges, which allows the particles to get close to each other sufficiently for making interactions via van der Waals forces [START_REF] Akbal | Copper, chromium and nickel removal from metal plating wastewater by electrocoagulation[END_REF]. The coagulants simultaneously form a sludge blanket to trap the suspended microplastics in the wastewater sample. The results show a removal efficiency higher than 90% for all experiments using proposing electrocoagulation. The highest removal efficiency, 99.24%, was obtained with pH of 7.5 and NaCl concentration of 0-2 g/L. Furthermore, the lowest tested current density of 11 A/m 2 , which is the best in view of energy consumption, was the most efficient for achieving the highest removal rate. [START_REF] Herbort | Alkoxysilyl induced agglomeration: a new approach for the sustainable removal of microplastic from aquatic systems[END_REF] suggested agglomeration based on alkoxy-silyl bond formation via sol-gel reactions as a new sustainable removal approach for treatment for the microplastics originating from inert products of the textile and cosmetic industries. Functionalized molecular precursors were initially synthesized in an inert atmosphere, then used for bio-inspired alkoxy-silyl formation. Meanwhile, microplastics adhered together to form large three-dimensional agglomerates which can, afterward, be removed using costefficient filtration methods (Fig. 7). The sol-gel formed in this way is similar to hybrid organic-inorganic silica gels with a large variety of benefits and usages in sensors and optical materials, medicine, and corrosion protection [START_REF] Nicole | Advanced selective optical sensors based on periodically organized mesoporous hybrid silica thin films[END_REF]).

The mechanisms of degradation of microplastics are not fully known. [START_REF] Brandon | Long-term aging and degradation of microplastic particles: comparing in situ oceanic and experimental weathering patterns[END_REF] studied the degradative changes of the chemical structure of two types of microplastics including polypropylene and polyethylene for 3 years under simulated realistic weather conditions. According to FTIR analysis, they found some slight nonlinear changes with time in moieties like carbonyl, hydroxyl and carbon-oxygen bonds, implying microplastics slow degradation.

The exposure of macroplastics to elements [START_REF] Colom | Natural and artificial aging of polypropylene-polyethylene copolymers[END_REF][START_REF] Gulmine | Degradation profile of polyethylene after artificial accelerated weathering[END_REF], microorganisms [START_REF] Pathak | Review on the current status of polymer degradation: a microbial approach[END_REF], catalysts [START_REF] Hazrat | A study on thermo-catalytic degradation for production of clean transport fuel and reducing plastic wastes[END_REF], and photo-active materials [START_REF] Li | Photocatalytic degradation of polyethylene plastic with polypyrrole/TiO 2 nanocomposite as photocatalyst[END_REF]) has been widely reported but there are no large number of publications on microplastics degradation. [START_REF] Liu | New insights into the aging behavior of microplastics accelerated by advanced oxidation processes[END_REF] studied the long-term aging behavior of polystyrene and polyethylene microplastics treated by a combined heat-activated persulfate and Fenton method in the aquatic environment. According to their results, O/C ratio and the average size of the microplastics play a role in the adsorption capacity and surface properties, which affect the microplastic oxidation rate significantly. Degradation of polyethylene microplastics in artificial seawater under dark and UV radiation to evaluate the structural and morphological alterations has been recently reported [START_REF] Costa | Degradation of polyethylene microplastics in seawater: insights into the environmental degradation of polymers[END_REF][START_REF] Pathak | Review on the current status of polymer degradation: a microbial approach[END_REF]. FTIR analysis of the initial materials and products (Fig. 8) evidenced the stronger degradative role of the artificial seawater compared to UV illumination. This was confirmed by an increase of the organic content of the medium. As a control experiment, the researchers exposed microplastics to UV light alone for the same period of time and did not observe significant changes in the chemical structure, thus showing implying that salt is needed to form oxidized sites. SEM images also revealed the effect of salt on microplastic surface morphology, in which observable crack lines were clearly detected. This strengthens the fact that media salinity facilitates microplastic degradation [START_REF] Vasile | Practical guide to polyethylene[END_REF].

Photocatalytic degradation of the organic contaminants is a destructive process triggered by absorbed photons on the surface of a semiconducting material for generating hyperactive radicals. Depending upon the semiconductor band gap, the energy of the incident photons and, subsequently, the suitable light source must be chosen. Photooxidation degradation, that is oxidation in the presence of light source and air, is the mostly common process for the breaking of polymeric chains [START_REF] Yousif | Photodegradation and photostabilization of polymers, especially polystyrene[END_REF]. Different mechanisms were suggested for the photochemical processes in the aqueous phase, in which hydroxyl radicals strongly promote the degradation reactions [START_REF] Li | Photocatalytic degradation of polyethylene plastic with polypyrrole/TiO 2 nanocomposite as photocatalyst[END_REF]. Heterogeneous photocatalytic degradation of low-density polyethylene microplastics has been recently investigated in the aquatic media over the rod-like ZnO nanoparticles [START_REF] Tofa | Visible light photocatalytic degradation of microplastic residues with zinc oxide nanorods[END_REF]. From the optical images, morphological changes including the appearance of wrinkles, brittleness, cracks and spots on photo-exposed surfaces of the microplastics were observed (Fig. 9). Also, the results revealed variations in elasticity properties of the sample exposed to photocatalytic conditions in comparison with nonirradiated wastewater, and this directly is in correlation to the changes in the strength of chemical bonds. Accordingly, the nanorods surface area influenced the stiffness degree of the microplastics and the photocatalytic performance. [START_REF] Liang | High photocatalytic degradation activity of polyethylene containing polyacrylamide grafted TiO 2[END_REF] obtained FTIR data which confirmed the presence of newly formed functional groups such as carbonyl and vinyl during the photocatalytic treatment, they proposed a mechanism for the mineralization of microplastics in wastewater (Eqs. 6-13):

(6) (-CH 2 -CH 2 -) n + OH ⋅ → (-CH 2 -CH ⋅ -) n + H 2 O (7) (-CH 2 -CH ⋅ -) n + O 2 → (-CH 2 -HCOO ⋅ -CH 2 -) n
Fig. 9 Decrease of the elasticity of microplastics after photocatalytic treatment with: A: ZnO (10 mM, 5 h), B: ZnO (3 mM, 5 h) and C: control. Optical images of the treated microplastics, i photo-irradiated, only, for 175 h, ii photo-irradiated in the presence of the ZnO nanorods, 3 mM for 5 h, iii photoirradiated in the presence of the ZnO nanorods, 10 mM for 5 h, iv photo-irradiated in the presence of the ZnO nanorods, 20 mM for 5 h, representing the development of holes, cracks, and spots [START_REF] Tofa | Visible light photocatalytic degradation of microplastic residues with zinc oxide nanorods[END_REF]) [START_REF] Sivan | New perspectives in plastic biodegradation[END_REF]

(-CH 2 -HCOO ⋅ -CH 2 -) n + (-CH 2 -CH 2 -) n → (-CH 2 -HCOOH -CH 2 -) n + (-CH 2 -CH ⋅ -) n (9) (-CH 2 -HCOOH -CH 2 -) n → (-CH 2 -HCO ⋅ -CH 2 -) n + OH ⋅ (10) (-CH 2 -HCO ⋅ -CH 2 -) n → carbonyl groups (norrish type I) (11) (-CH 2 -CO -CH 2 -) → (-CH 2 -CO ⋅ ) n + ( ⋅ CH 2 -CH 2 -) n (norrish type II) (12) (-CH 2 -CO -CH 2 -) → (-CH 2 -CO) n + (CH 2 = CH 2 ) n (13) (-CH 2 -HCO) + (-CH 2 -COOH) + (-CH 2 -CO -CH 2 -) → CO 2 + H 2 O
The design and degradative capability of the TiO 2 -based micro-and nanodevices for the photocatalytic treatment for microplastics have been studied [START_REF] Sekino | Fundamental study on nanoremoval processing method for microplastic structures using photocatalyzed oxidation[END_REF]. Holding high potential for water purification, remediation of environmental pollutions and medicinal usages, photo-active micromotors have attracted huge attention [START_REF] Eskandarloo | Light-harvesting synthetic nano-and micromotors: a review[END_REF][START_REF] Zhang | Microplastic pollution in China's inland water systems: a review of findings, methods, characteristics, effects, and management[END_REF]. [START_REF] Wang | Photocatalytic TiO 2 micromotors for removal of microplastics and suspended matter[END_REF] reported the use of a Au-decorated TiO 2 -based micromotor for the efficient photocatalytic removal of polystyrene microplastics from wastewater under UV illumination (Fig. 10). The micromotor propulsion is supplied by photochemical reactions in water and hydrogen peroxide initiated by electron-hole generation processes. Although the lack of selectivity was the most important disadvantage of this system. The need for low concentrations of H 2 O 2 for promoting the phoretic interactions, which enable the micromotor to move, makes this system actually non applicable in real wastewater.

Recently, Ariza-Tarazona et al. ( 2019) reported the ability of a protein-based N-TiO 2 photocatalyst to degrade the polyethylene microplastics in solid and aqueous phases. According to their results, photocatalytic treatment for the pollution resulted in 1.1% of the mass loss in the solid phase with the rate constant of 12.2 × 10 -4 /h after 20 h of treatment time. These values were reported to be 6.4% and 38.2 × 10 -4 /h in the aqueous phase. They demonstrated that surface area of the semiconductor and interactions between the photocatalyst surface and microplastics significantly influenced the removal efficiency. The SEM images and FTIR spectra have also provided good evidences to confirm the structural changes during the mineralization process.

To conclude, coagulation and agglomeration methods have been investigated to remove microplastics in the laboratory. The efficiency depends on the type of coagulant, pH, the chemical composition of the media and the concentrations. Although salinity of the media and the presence of illumination source are reported to affect modify the degradation of microplastic structures in the aquatic environment, degradative processes have been rarely studied. Of special interest are photocatalytic nanomaterials which make use of both irradiation and catalytic degradation.

Biological removal and ingestion

Microplastic removal by marine organisms [START_REF] Harrison | Interactions between microorganisms and marine microplastics: a call for research[END_REF] published "Interactions Between Microorganisms and Marine Microplastics: A Call for Research" to highlight the outstanding potential of microorganisms, including picoeukaryotes, bacteria and archaea, to facilitate the biological degradation of microplastics in coastal sediments. They discuss the interactions between synthetic microplastics and marine microorganisms. Many reports present the biological degradation of natural and Fig. 10 Photocatalytic removal of polystyrene microplastics in water by a Au-Ni-TiO 2 micromotor [START_REF] Wang | Photocatalytic TiO 2 micromotors for removal of microplastics and suspended matter[END_REF]. The microscopy images represent the efficient interactions leading to photodegradation reaction of the microplastics on the surface of the micromotor synthetic macroplastics [START_REF] Ahmed | Biodegradation of plastics: current scenario and future prospects for environmental safety[END_REF][START_REF] Bettas Ardisson | Biodegradation of plastics in soil and effects on nitrification activity. A laboratory approach[END_REF].

Polyethylene microplastics fragmentation and size alteration ingested by Antarctic Krill (Euphausiasuperba), a planktonic crustacean, were studied by a group of environmentalists in Australia [START_REF] Dawson | Turning microplastics into nanoplastics through digestive fragmentation by Antarctic krill[END_REF]. The mechanism of fragmentation and type of interactions between microplastics and zooplankton, in which biota-facilitated degradation occurs, still remains unclear but the experiments confirmed that smaller microplastics are much easily fragmented under environmental conditions [START_REF] Ter Halle | Understanding the fragmentation pattern of marine plastic debris[END_REF]. Electron microscopic evidences confirmed the shrinking of polyethylene microplastics through fragmentation, where the physical size decreased from ~ 31 μm for the microplastics to less than 1 μm for the fragmentation products (Fig. 11) [START_REF] Dawson | Turning microplastics into nanoplastics through digestive fragmentation by Antarctic krill[END_REF]. The findings thus provide evidence for the biologically-mediated transformation of microplastics to nanoplastics.

The preliminary results of a four-month study on removal of high-density polyethylene secondary microplastics in seawater using two types of indigenous marine communities including Agios consortium and Souda consortium were published by [START_REF] Cocca | Proceedings of the international conference on microplastic pollution in the mediterranean sea[END_REF]. According to the recorded weight reduction results, the Souda consortium was more efficient. Interestingly, their results on monitoring the cell content and populations suggested that microplastics acted as a rich carbon source to feed the organisms. While the protein content of both communities decreased during the [START_REF] Dawson | Turning microplastics into nanoplastics through digestive fragmentation by Antarctic krill[END_REF] experiment, the increase in cell's carbohydrate content implies enhancement in microbial tendency to adhere to the microplastics surfaces, which is the first step in microbial polymer degradation [START_REF] Sivan | New perspectives in plastic biodegradation[END_REF]).

An important concern about the microplastics is their small size and negligible weight which enables them to easily spread over the long distances by wind-driven free ocean surface layer movements [START_REF] Ivleva | Microplastic in aquatic ecosystems[END_REF]. Studies have also indicated that microplastics in sediments are normally in the range of 0-3146 particles/kg of dry weight sediment, which highlights the urgent need of designing comprehensive experimental routes for the removal from aquatic ecosystems (Maes et al. 2017). The growth of microbial assemblages on the surface of microplastics should form a new platform prone to settle other organisms like microscopic fungi and microalgae. Although these microorganism communities can probably catalyze the metabolic reactions of microplastics resulting in their further biodegradation, the transport of nonnative organisms, which does not naturally occur in such environments, can induce negative effects on marine ecosystems variety [START_REF] Urbanek | Degradation of plastics and plastic-degrading bacteria in cold marine habitats[END_REF]. [START_REF] Paço | Biodegradation of polyethylene microplastics by the marine fungus Zalerion maritimum[END_REF] explored the capability of fungus Zalerion maritimum, a naturally occurring fungus in marine ecosystems, for the polyethylene microplastics biodegradation based on mass and size variations of the microplastics in a batch reactor (Fig. 12). They measured the concentration of the contaminant at various time intervals to recognize the reaction rate order and found that biodegradation kinetics of the microplastics obeys to fractional order which can be described in terms of the complex mechanism expected for such reactions [START_REF] Boudart | Kinetics of heterogeneous catalytic reactions[END_REF].

The evidences from biological content measurements, such as protein reduction and carbohydrate increase with exposure time, revealed that the Z. maritimum community possibly uses the microplastics as a nutrient source, an identical finding described earlier [START_REF] Sivan | New perspectives in plastic biodegradation[END_REF]. Noteworthily, the results of the electron and optical microscopy clearly revealed that the biological compounds were present on the surface of the microplastic pellets, typically, highlighting the aforementioned potential of such organisms for triggering the effective biological degradation of microplastics from their surfaces. From FTIR analysis, the increases in the new bands' intensity appearing at regions 3700-3000, 1700-1500 cm -1 , and 1200-950 cm -1 , are, respectively, attributed to the hydroperoxide and hydroxyl groups, carbonyl groups, and double bonds, implying the oxidative degradation of polyethylene microplastics. Such variations were already observed in another work in which low-density polyethylene microplastics were photocatalytically removed from wastewater under visible light [START_REF] Tofa | Visible light photocatalytic degradation of microplastic residues with zinc oxide nanorods[END_REF]. [START_REF] Auta | Screening of Bacillus strains isolated from mangrove ecosystems in Peninsular Malaysia for microplastic degradation[END_REF] investigated the removal of different microplastics composed of polyethylene, polystyrene, polyethylene terephthalate and polypropylene by Bacillus cereus and Bacillus gottheilii, two types of Bacillus bacterial strains isolated from mangrove sediments (Fig. 13). In addition to scanning the morphological and structural changes using electron microscopy and FTIR analyses, the rate of biodegradation was assessed via measuring the microplastics weight loss. The fastest mass reduction (0.0019/ day) and shortest degradation half-life (363.16 days) was found using B. cereus isolate and polystyrene microplastics, whereas B. gottheilii on polyethylene gave 0.0016/ day and 431.25 days. Microplastics biodegradation can also be monitored performed by scanning electron microscopy (SEM), where several cracks, holes and erosions were observed [START_REF] Sowmya | Biodegradation of polyethylene by Bacillus cereus[END_REF]. By comparing the degradation results, bond cleavages and chemical alterations scanned by FTIR analysis, B. gottheilii appeared as a better potential microplastic degrader.

Microplastic removal by bacteria

Microplastic ingestion

The number of papers discussing microplastics ingestion by organisms as a removal strategy is limited [START_REF] Arossa | Microplastic removal by red sea giant clam (Tridacna maxima)[END_REF][START_REF] Cole | Microplastic ingestion by zooplankton[END_REF][START_REF] Hall | Microplastic ingestion by scleractinian corals[END_REF]. Based on the results of coherent anti-Stokes Raman scattering and fluorescence microscopy, [START_REF] Cole | Microplastic ingestion by zooplankton[END_REF] suggested that zooplankton's function and health can be negatively affected by micro-sized polystyrene plastic debris. They also stated that microplastics adherence to the external carapace of zooplankton considerably influenced the organism feeding, a crucial hindrance factor against normal growth conditions in marine wildlife [START_REF] Van Franeker | Monitoring plastic ingestion by the northern fulmar Fulmarus glacialis in the North Sea[END_REF]. Although this work was not aimed to introduce a way for microplastics removal, the results reveal the high capacity of zooplankton to remove 1.7-30.6 μm polystyrene microplastics through uptake and ingestion inside the organism body, in which their intestinal tracts can hold the microplastics up to 7 days after entrance. [START_REF] Hall | Microplastic ingestion by scleractinian corals[END_REF] also reported capture and ingestion of polypropylene microplastics by scleractinian corals with a consumption rate of 50 μg plastic cm -2 /h, and evidenced the presence of microplastics in mesenterial tissue within gut cavity of the coral for at least 24 h. However, the mechanism of the affection which potentially interferes with coral growth, physiology, and survival still remains unclear. This research was important since inshore coral reefs ecosystem, the coastal areas, is easily available as touristic regions for long periods and many of water treatment plants also daily release their effluents containing microplastics debris, as mentioned above, to such environments.

The evaluation of the capability of the Red Sea giant clam, Tridacna maxima, for the removal of 53-500 μm polyethylene microplastics from wastewater (Fig. 14) has been reported by [START_REF] Arossa | Microplastic removal by red sea giant clam (Tridacna maxima)[END_REF]. Their results revealed the key role of clam's shells to remove the microplastics by sorption on the surface, as they resulted in 66.03% of removal from the wastewater. The authors suggested that this marine biota can be used for rapid removal of the anthropogenic microplastics debris from aquatic surfaces in the regions with low concentration of microplastics, a fact partly demonstrated by previous researchers [START_REF] Martí | Low abundance of plastic fragments in the surface waters of the Red Sea[END_REF]. The effect of microplastics initial concentration and clam's size on the removal rates through active, microplastics retention inside the body, and/or passive, trapping by the shells, pathways was investigated. The results indicated that big clams tend to retain bigger microplastics inside their bodies, while small clams did not show any sharp preference for them. Moreover, at any concentration, the microplastics easily penetrated into the digestive system of all clams used in this experiment. According to their measurements, the passive route of removal accounts averagely for 35.95% of the microplastics, leading to a significant increase in the density of microplastics attached to clam's shells and normally, the larger clams adsorb higher concentrations of the microplastics. As a whole, this research opened a new window for environmental pollution management making use of the marine biota possessing complex surface and structure. But, finding the mechanism of interactions between the microplastics and organisms needs to be explored. Accordingly, we will be able to opt the type of sorbent biota and operational conditions for achieving the highest microplastics removal efficiency.

In brief, microplastics can provide a platform to grow a variety of microbial assemblages. Biological degradation using microorganisms, e.g., zooplankton, marine fungi and bacterial strains, has been confirmed to be suitable to remove microplastics at low concentrations. The mechanism of interactions and fragmentation is not well understood. Microplastics, in the most of case studies, acted as a nutrient source for the organisms. This could not convince us to consider "ingestion" as a removal strategy to treat the microplastic pollutions.

A comparison of treatment methods for removal of the microplastics from the environment is represented in Table 3. The table briefly summarizes the advantages, disadvantages, and the efficiency of the projects utilized for the last few years. It is obvious that membrane-included technologies are still the most reliable techniques to remove microplastics in the practical applications. However, it seems that finding the suitable alternatives to couple with these methodologies is the main goal of the ongoing projects.

Conclusion

After short surveying on sources, occurrence, and transport pathways of microplastics, recent studies on three major approaches including chemical, biological, and physical methods to remove them from the environments have been summarized. Sorption and filtration processes coupled with membrane bioreactors lead to removing a high percentage of microplastics in the influents entering into the water treatment plants but these systems acted themselves as daily microplastic sources since the effluents are directly released to aquatic environments. Conventional activated sludge strategy is also used to treat wastewaters in water treatment plants but shows lower efficiency than membrane bioreactor technologies which makes it a less popular methodology. Electrocoagulation and agglomeration have also become reliable techniques for easy separation of microplastics but are needed to be coupled with eminent extra filtration stages. Photocatalytic degradation using TiO 2 and ZnO semiconductors is experimentally revealed to be a suitable method among the aforementioned strategies. FTIR and electron microscopy analyses are widely served to elucidate any structural alterations during the degradation process. Biological degradation has been investigated over the bacterial strains and marine organisms. Appearance of hydroperoxide and hydroxyl groups, carbonyl groups, and double-bond characteristic bands in the FTIR patterns of the treated microplastics suggests the oxidative mechanism for [START_REF] Hall | Microplastic ingestion by scleractinian corals[END_REF] and [START_REF] Paço | Biodegradation of polyethylene microplastics by the marine fungus Zalerion maritimum[END_REF] biological removal. Ingestion by organisms has still not considered as a removal strategy but Red Sea giant clam exhibits good potential for microplastics sorption on the shells besides its degradation in digestive system.

Fig. 1

 1 Fig. 1 Microplastics sources, transformation and transport

Fig. 2

 2 Fig. 2 Dynamic membrane experimental setup and graph showing the decrease in turbidity with time when microplastics are removed (Li et al. 2018b). TMP transmembrane pressure, NTU nephelometric turbidity unit

Fig. 3

 3 Fig.3Microplastics removal efficiency, in terms of number of microplastics per liter, MP/L, after treatment by final-stage treatment technologies(Talvitie et al. 2017a)

Fig. 5 Fig. 6

 56 Fig.5Removal of polyethylene microplastics from the wastewater using Fe 3+ coagulation, sedimentation and ultrafiltration processes[START_REF] Ma | Removal characteristics of microplastics by Fe-based coagulants during drinking water treatment[END_REF]. The effect of polyacrylamide (PAM), pH and the

Fig. 7 Fig. 8

 78 Fig.7Removal of polyethylene microplastics using alkoxy-silylinduced agglomeration. The microscopy images of the microplastic before agglomeration (above left) and after that (above right). The

Fig. 11

 11 Fig. 11 Fate of microplastics ingested by Antarctic krill. Microplastics on a filter paper isolated from digested krill with auto-fluorescent mandible (a), digestive gland tissue (b), midgut and digestive gland tissue (c), mandible with the fragments embedded into the surface (d), and fecal pellet with microplastics under bright-field and fluores-

Fig. 12

 12 Fig. 12 Structural and morphological changes of the treated microplastics through biodegradation over fungus Z. maritimum in a batch reactor model (Paço et al. 2017). a NMR spectra, b FTIR patterns, c optical images and d electron microscopy images of the treated microplastics after 28 days. NMR: nuclear magnetic resonance. FTIR: Fourier transform infrared

Fig. 14 Fig. 13

 1413 Fig. 14 Polyethylene microplastics removal from wastewater over Red Sea clams (Arossa et al. 2019). Experiments involved four concentrations of microplastics over two different sizes of T. Maxima

Table 1

 1 Persistency criteria for contaminants in different media according to the REACH Annex XIII[START_REF] Verschoor | Towards a definition of microplastics: Considerations for the specification of physico-chemical properties[END_REF]

	Compartment	Half-life (days)
	Marine water	> 60
	Fresh or estuarine water	> 40
	Marine sediment	> 180
	Fresh or estuarine sediment	> 120
	Soil	> 120

Table 2

 2 Average microplastic concentrations before and after treatment with different technologiesData are given in number of microplastics per liter of effluent a Pore size in μm(Talvitie et al. 2017a)

	Treatment	Effluent type	Before (MP/L)	After (MP/L)	Removal (%)
	Disk filter 10 a	Secondary	0.5	0.3	40.0
	Disk filter 20 a	Secondary	2.0	0.03	98.5
	Rapid sand filter	Secondary	0.7	0.02	97.1
	Dissolved air flotation	Secondary	2.0	0.1	95.0
	Membrane bioreactor	Primary	6.9	0.005	99.9

Table 3

 3 Comparison of different methods for the removal of microplastics from the environment Aggregation of the microbial assemblages on the surface, environmental conditions cannot be easily controlled, difficulties to analysis of the products in large scale, lack of reproducibility, finding the suitable microbial community Arossa et al. (2019), Auta et al. (2017), Cole et al. (2013), Dawson et al. (2018),

	References	Lagarde et al. (2016) and Zhang et al.	(2017)		Li et al. (2018b)			Lares et al. (2018)		Gurung et al. (2016)				Leslie et al. (2017), Long et al. (2019) and	Mintenig et al. (2017)			Herbort et al. (2018) and Ma et al. (2019)			Perren et al. (2018)					Sekino et al. (2012), Tofa et al. (2019),	Wang et al. (2019) and Ariza-Tarazona	et al. (2019)
	Drawbacks	Nonrecyclable method, chemical adher-	ence of the microplastics to the surface	and poisoning it	much cleaning to avoid excessive mem-	brane fouling, energy demand, sludge	accumulation due to its flat structure	Shape dependency of the removal per-	centage, membrane fouling	long retention times in the tank, the large	sedimentation surface, the high cost of	energy and the processing and disposal	of sludge	Act as secondary microplastic sources,	disability to treat small-sized micro-	plastics, sludge aggregation, large	mechanical devices	Addition of chemicals to media, nonus-	able for large microplastics,		Repeated need of replacing the sacrificial	anode, cathode passivation, nonusable	in areas without electricity			Lack of selectivity, lower efficiency com-	pared to the other methods, difficult to	scale up, generation of secondary-type	organic pollutants, huge photoreactors	setup, energy consuming, photocatalyst	recovery
	Advantages	High affinity of the cut surfaces to sorb	tiny microplastic particles, selectivity	based on microplastics surface charge	Low filtration resistance, low trans-	membrane pressure, easy operation,	nonchemical treatment	Using combined advanced treatment	methods with porous membranes,	Robust, cost-effective, flexible, treating a	wide range of influent concentrations,	applicable for large-scale treatments		Efficient mix of sorption-biological treat-	ment processes, low maintenance costs,	simple operation		Suitable to remove small microparticles,	controllable operational conditions,	simple mechanical devices	No chance of secondary pollution,	suitable for the removal of smallest	particles, sludge minimization, energy	efficiency, cost-effectiveness, flexibility	to automation	No need to add other chemicals, extend-	able applications to solar light as a	renewable and pollution-free energy	source, efficient mineralization, envi-	ronmentally friendly remediation	Simplicity and safety for large-scale use,	low operating costs, practically applica-	ble in different environments, flexibility	to handle a wide range of wastewater	characteristics and flows
	Method Efficiency%	Adsorption on green microalgae 94.5			Dynamic membranes > 90			Membrane bioreactors > 99		Conventional activated sludge 98				Wastewater treatment plants > 95				Classic coagulation and 61	agglomeration methods		Electrocoagulation > 90					Photocatalytic degradation					Biological degradation Depends on type of	microbial community,	mostly > 20

Acknowledgements This work is financially supported by SAFEA of China (High-End Foreign Expert Project # G20190241013) and the scientific research startup fund of Shaanxi University of Science and Technology.


L Cattin

M A Cherif

Z El Jouad

H Ftouhi

E M El-Menyawy

A El Mahlali

S Touihri

L Arzel

M Addou

P Torchio

Jean Christian Bernède

On the contribution of fullerene to the current of planar heterojunction organic solar cells

Keywords: Organic solar cells, Planar heterojunction, External Quantum Efficiency, Absorption, Optical Modeling

Recently, significant progress in the field of organic photovoltaic cells was obtained by substituting new electron acceptor molecules to the fullerene, which was attributed to the fact that the fullerene absorption is quite small. Nevertheless, we demonstrate in the present work that, in the case of inverted cells, i.e. when the transparent bottom electrode is used as cathode, the contribution of fullerene to the Jsc short-circuit current of the cells, if not dominant, is not negligible; and that mainly in the short wavelength spectral range. The experimental results are confirmed by an optical simulation. Due to this significant contribution to Jsc, the light transmission of the transparent electrode towards the UV-part of the spectrum is crucial for inverse cell performances. When a transparent conductive electrode based on an alternative dielectric/metal/dielectric structure is substituted to ITO, such as ZnS/Ag/TiO 2 , the study allows to obtain promising results, although there is a loss of performance due to the decrease of transmission of TiO 2 below 400 nm in wavelength.

Introduction

Due to their lightweight, flexibility, semi-transparency [START_REF] Christopher | Organic photovoltaic greenhouses: a unique application for semi-transparent PV?[END_REF] organic photovoltaic cells (OPVs) are complementary to silicon cells. The spectacular progress achieved in the field of OPVs is due to the use of low energy gap molecules [START_REF] Liu | Low bandgap semiconducting polymers for polymeric photovoltaics[END_REF], new electron acceptors [START_REF] Zhang | Non fullerene Acceptor Molecules for Bulk Heterojunction Organic Solar Cells[END_REF], and optimization of their geometry [START_REF] Ameri | Organic tandem solar cells: A review[END_REF][START_REF] Cattin | Smart geometrical approach to intercalate a highly absorbing and quite resistive electron donor layer in ternary organic photovoltaic cells[END_REF]. A basic OPV is made up of an organic active layer, consisting of an electron donor (ED) and an electron acceptor (EA), sandwiched between two electrodes, one of which being transparent. It is usual to employ this transparent layer, generally an ITO layer, as an anode. Such a configuration is called classic OPV (COPV). In contrast, when the transparent conductive electrode (TCE) is used as a cathode it is called inverted OPV (IOPV).

It was shown that IOPVs are more stable than COPVs, hence the interest in studying these inverted structures, even if, at times, their performances are inferior to that of COPVs [START_REF] Zhang | Recent development of the inverted configuration organic solar cells Solar Energy[END_REF], but this is not sysyematically the case when the ED is SubPc [START_REF] Bernède | Organic photovoltaic cells: History, principle and techniques[END_REF].

Whatever their geometry, IOPV or COPV, two configurations are possible for the organic layer, either bulk-heterojunction (BHJ) or planar-heterojunction (PHJ). If BHJ-OPVs allow achieving highest efficiencies, PHJ-OPVs, due to their simplicity, permit obtaining more reproducible results [START_REF] Bernède | Organic photovoltaic cells: History, principle and techniques[END_REF]. Moreover, while when a fullerene derivative is used as electron acceptor in BHJ-OPV it is well admitted that its contribution to the OPV current is small, it appears that it is not so evident in the case of inverted PHJ-OPVs (i.e. PHJ-IOPV). As a matter of fact, in the present work, using SubPc/C 60 as heterojunction ED/EA in PHJ-IOPVs, we show that, if the SubPc contribution to the external quantum efficiency (EQE), is dominant, that of C 60 is far to be negligible. This experimental result is comforted by optical simulation of the OPVs. On the other hand, it is well admitted that flexible optoelectronic is promised rapid development in the coming years [START_REF] Li | Flexible and Semitransparent Organic Solar Cells[END_REF]. Therefore, ITO being poorly flexible, it is necessary to look for new flexible transparent electrodes. Among the possible substituents, Dielectric/Metal/Dielectric (DMD) structures are one of the most promising [START_REF] Cattin | Toward indium-free optoelectronic devices: Dielectric / Metal / Dielectric alternative conductive transparent electrode in organic photovoltaic cells[END_REF]. We have probed two of these DMD structures as cathodes in the inverted OPVs studied in the present work. We show that the transmission spectrum of the TCEs is primordial for OPVs performances.

Description of approach and techniques

As described in supporting S1, OPVs were realized by deposition under vacuum on TCE coated glass substrates [START_REF] Cattin | Improved performance of organic solar cells by growth optimization of CuI/MoO 3 double anode buffer[END_REF]. The optically and electrically active core of the cells consisted of the planar heterojunction SubPc/C 60 . The transparent electrode was either ITO or a DMD structure. Two kinds of DMD structure nm/10nm/31nm for the different layers respectively [START_REF] Bernède | Dielectric/Metal/Dielectric flexible transparent electrodes, from smart window to semi-transparent solar cells[END_REF].The second electrode was an aluminum layer thick of 100 nm. that to optimize the charge collection it is necessary to introduce buffer layers at the electrode/organic material interfaces thin MoO 3 layer as hole transporting layer (HTL), MoO hole collector, even if the work function of the anode itself is not ver transporting layer (ETL), Alq between the cathode and the fullerene. kinds of DMD structures were used, ZnS/Ag/TiO 2 , with the for the different layers [START_REF] Cherif | Tailored ZnS/Ag/TiOx Transparent and Conductive Electrode for Organic Solar Cells[END_REF], and MoO 3 /Ag/MoO 3 , with 35

The second electrode was an aluminum layer thick of 100 nm. that to optimize the charge collection it is necessary to introduce buffer layers at the interfaces [START_REF] Godoy | Effects of the buffer layer inserted between the transparent conductive oxide anode and the organic electron donor[END_REF]. At the interface anode/SubPc we have introduced a layer as hole transporting layer (HTL), MoO 3 being well known as very efficient hole collector, even if the work function of the anode itself is not very high [ Alq 3 aluminum between the cathode and the fullerene. When used in COPVs, the Alq 3 ETL was proved to be very efficient [START_REF] Song | Small-molecule organic solar cells with improved stability[END_REF]. All chemical products were The second electrode was an aluminum layer thick of 100 nm. It is known that to optimize the charge collection it is necessary to introduce buffer layers at the . At the interface anode/SubPc we have introduced a being well known as very efficient y high [START_REF] Sun | Efficient inverted polymer solar cells with thermal-evaporated and solution-processed small molecular electron extraction layer[END_REF]. As electron um), was introduced ETL is called exciton All chemical products were he different layers evocated above were sequentially As evocated before, the main goal of this work is to study the contribution of C 60 to the photocurrent of OPVs and, more precisely, to compare its contribution in the classical OPVs to that in the inverted OPVs. To carry out this comparison, all other things being equal, we used in the inverse cells, the same materials as in the classical cells. Therefore, knowing that for classical cells the configuration was Glass/ITO/MoO 3 /SubPc/C 60 /Alq 3 /Al, for reverse structures we used the configuration Glass/ITO/Alq 3 /C 60 /SubPc/MoO 3 /Al. We have already shown that, in the case of inverted OPVs it was difficult to achieve reproducible performances [START_REF] Mohammed-Krarroubik | The Influence of Deposition Rates on Properties of AlPcCl Thin Films and on the Performance of Planar Organic Solar Cells[END_REF]. Actually, it was demonstrated that the open circuit voltage (Voc), the fill factor (FF) and therefore the OPV efficiency () depend on the morphology of the SubPc layer, which itself depends on its deposition rate. Optimum results were obtained when the deposition rate of SubPc was 0.02±0.01nm/s, the SubPc films being homogeneous. For higher deposition rate, the Voc and FF values and the reproducibility of the results decrease. It was also shown that the optimum thickness of the SubPc layer was 16 nm, for smaller or higher thickness the OPV efficiency decreases [START_REF] Mohammed-Krarroubik | The Influence of Deposition Rates on Properties of AlPcCl Thin Films and on the Performance of Planar Organic Solar Cells[END_REF]. At the beginning of the present work, even after optimization of the deposition rate and film thickness of SubPc, some variations in Voc and FF values persist.

Actually, the SubPc purity given by our supplier was 85 at. %, which means that from one batch to another one some significant differences can be encountered. This crucial point should be overcome. Therefore we had to purify the material. It was shown earlier that it was possible to purify phthalocyanine dyes by simply carrying out successive sublimations. Indeed, by using the same charge in the sublimation crucible, there is a progressive "auto purification" of the phthalocyanine dye [17]. Taking advantage of our experience acquired with the CuPc [START_REF] Berredjem | Optimisation of the interface "organic material/aluminium"of CuPc/C60 based photovoltaic cells[END_REF] we undertaken the purification of SubPc by a series of initial deposits which aimed to optimize the efficiency of the cells. Typical results are summarized in Figure 2. It can be seen that the performances of the OPVs depend very significantly of the number of deposits made, mainly through the values of Voc and FF. It is clear that Voc and FF increase with the number of cycles of deposits. For the first deposition, SubPc layer quality is very poor as shown by the J-V characteristics obtained.

Then, up 10 cycles, there is a progressive improvement of these characteristics due to the increase of Voc and FF while Jsc does not vary significantly after 3 cycles. From 10 cycles the characteristics are optimal and stable. It must be noted that our experience with CuPc discard the hypothesis of a possible artifact due to a problem on our equipment to deposit the film. Actually when the purity of CuPc is 99%, it is not necessary to carry out successive sublimations to obtain stable results, while it is necessary when the purity of CuPc is only 95% [START_REF] Berredjem | Optimisation of the interface "organic material/aluminium"of CuPc/C60 based photovoltaic cells[END_REF]. The effectiveness of our process in purifying these phthalocyanine dyes shows that some impurities present in the commercial powder have smaller sublimation temperature than the dyes, which allows the purification through deposition cycles.

After stabilization of the OPCs performances, i.e. purification of the SubPc present in the sublimation crucible, we proceeded to the comparison of the J-V characteristics and the EQE spectra of inverted and classical OPVs. The J-V characteristics of the inverted and classical OPVs manufactured during the present work are represented in figure 3. The champion IOPV produced a Jsc of 6.12 mA/cm 2 , a Voc of 0.99 V, a FF of 0.61and η of 3.76%, while Rs = 20 Ω and Rsh = 1800 Ω. On the other hand, the COPV produced a Jsc of 7.65 mA/cm 2 , a Voc of 1.00 V, a FF of 0.58 and η of 4.44%, while Rs = 15Ω and Rsh = 1300 Ω . The obtained results are rather close, the main difference being due to Jsc, which is smaller in the case of IOPV, as it is often the case in PHJ-OPVs. Typical EQE spectra of both types of OPVs are presented in Figure 4 and absorption spectra are shown in Figure 5. In the EQE spectra the signal situated between 500 nm and 650 nm can be attributed to SubPc. In the short wavelengths spectral range, the signal below 400 nm corresponds to the second absorption band of SubPc, but above all to C 60 . As shown by the absorption curves, the intermediate domain between these two contributions must be mainly attributed to C 60 . It is remarkable to note that the C 60 contribution in EQE is far to be negligible in the case of IOPV. In order to check the importance of the contribution of C 60 to the short circuit current we have studied series of IOPVs with different thicknesses of C 60 : 20 nm, 30 nm and 40 nm. Typical J-V characteristics are reported in the supporting information S4. The main information, which confirms the importance of the contribution of C 60 , is that the value of Jsc decreases when the thickness of C 60 decreases: Jsc = 4.10 mA/cm 2 when the 7 thickness of C 60 is 20 nm, Jsc = 4.95 mA/cm 2 for 30 nm and Jsc = 6.12 mA/cm 2 for 40 nm of C 60 . Beyond, Jsc starts to decrease due to the increase of the series resistance.

It can be seen in Figure 5 that the overall shape of the absorption spectra of IOPV and COPV are rather similar. Nevertheless, the ratio of the values measured at 450 nm and 590 nm (OD 450 /OD 590 nm) is different. It is greater than 1 in the case of IOPVs and smaller than 1 for COPVs, i.e. the contribution of C 60 is dominant in IOPVs. Moreover, it can be noted that the C 60 contribution in the UV spectral domain is very important.

Since the EQE spectra are clearly dependent of the absorption spectra of the organic layers, as shown by the Figures 3 and4

Optical simulation

A numerical model based on a Transfer Matrix Method (TMM), is employed to predict the optical behavior of the considered electrodes or organic solar cells [START_REF] Bou | Indium Tin Oxyde-free Transparent and Conductive Electrode based on SnOx|Ag|SnOx for Organic Solar Cells[END_REF][START_REF] Peres | ZnS/Ag/TiO 2 Multilayer Electrodes with Broadband Transparency for Thin Film Solar Cells[END_REF]. It supposes that the stacks are composed of flat, massive and homogeneous layers, that the air surrounds the multilayer, and that light penetrates at normal incidence. The input data of our simulation algorithm are the thicknesses of each material and their complex optical constants, previously obtained by ellipsometric spectroscopy or by scientific literature [START_REF] Palik | Handbook of Optical Constants of Solids[END_REF]22]. The parameters used for the simulation of transmission spectrum are given in the supporting information S5 and in the references [START_REF] Palik | Handbook of Optical Constants of Solids[END_REF][START_REF] Bou | Numerical and Experimental Investigation of Transparent and Conductive TiOx/Ag/TiOx Electrode[END_REF].

This method makes it possible to calculate the optical properties such as the reflectance R, the transmittance T and the absorbance A, and to optimize the thicknesses of our structures [22,[START_REF] Bou | Numerical and Experimental Investigation of Transparent and Conductive TiOx/Ag/TiOx Electrode[END_REF]. As an example, we are able to compare the modeled transmittance to those measured (figure 6) from a typical ZnS/Ag/TiO x electrode manufactured by a vacuum electron beam evaporator on a glass substrate, then characterized by UV-Visible spectrophotometer. Results are globally in good agreement for such Glass/ZnS (31 nm)/Ag (10 nm)/TiOx (31 nm) electrode (the measured curve is in full line while the simulated curve is in dashed line). However, only the absorbance in the photo-active layers is likely to produce photocurrent in the cells. Therefore, we calculated the useful intrinsic absorption in the only SubPc and C 60 active layers in order to correlate our results with the EQE ones in the same spectral domain (figure 8). It appears clearly in the classical design two absorption bands: one below the wavelength of 550 nm which is mainly due to the absorption of the C 60 acceptor material, and one above this wavelength that is mainly attributed to the SubPc donor molecule. In the inverted cell, the optical contribution of the C 60 is increased between 425 and 525 nm, while those of the SubPc is decreased between 525 and 600 nm. These variations can be explained by the optical interferences taking place in the whole stack, which modify the distribution of the electromagnetic field and thus the absorbance, and by the thickness of the SubPc that is a little smaller in the inverted cell. It is also remarkable to note some strong similarities with the EQE curves of the figure 4 as the presence of a "crossing" wavelength (around 500-525 nm) between 2 spectral domains where the COPV is better than IOPV (in the [500-520, 600]-nm range), and where the IOPV is better than COPV (in the [350-400, 500-520]-nm range). This demonstrates the significant influence of the absorption on the EQE of the OPVs, and the non-negligible contribution of the C 60 acceptor molecule in its absorption range.

Inverted OPVs with different transparent conductive electrodes

After the study of OPVs using ITO as TCE, we have substituted DMD to ITO, using two different types of DMD: ZnS/Ag/TiO 2 and MoO 3 /Ag/MoO 3 . We have used them as cathode, the main goal of the present study being not to match the performances obtained with ITO, but to study the influence of the transmission curves of the new TCEs on the electrical characteristics of IOPVs. The transmission curves of the different TCEs used are visible in Figure 9. It can be seen that there is a decrease of the light transmission of the ZnS/Ag/TiO 2 structure for wavelengths below 375 nm, while in the spectral range above 550 nm, the MAM structure is the TCE that transmits the least. In Figure 9 we have also introduced the absorption spectra of SubPc and C 60 in order to facilitate the discussion of the shape of the EQE curves further shown. If the main contribution of the absorption spectrum of SubPc is situated in the wavelength range that corresponds more or less to the maximum transmission range of the three TCEs, it is not the case with C 60 . Actually, the main contribution of the Typical J-V characteristics obtained with the new TCEs, when they are substituted to ITO in IOPVs, are shown in Figure 10 and summarized in Table 1, while the corresponding EQE spectra are shown in Figure 11. The performances are sensibly smaller than that achieved with the ITO electrode. Actually, in the case of ZnS/Ag/TiO 2 -TCE, we have already shown that the performances of the OPVs, are limited by the diffusion of some Ag towards the surface of the TCE [24]. In the case of MoO 3 /Ag/MoO 3 it can be seen that with the ETL used in the present work, Alq 3 , the result is very bad. As a matter of fact, it is known that the value of the work function (W F ) of MoO 3 is higher than 5 eV, which makes that it is a very efficient HTL [25] but reversely a very bad ETL. So, when used as cathode in IOPVs the efficiency of the cells is very small. Nevertheless, when a small W F material, Ca, is introduced at the interface cathode/organic material, a far better result is obtained. Knowing that the goal of the present study was not to overpass some record but to study the effect of the transmission spectra of the new TCE on the IOPVs performances we focused our attention on the EQE spectra. It can be seen in Figure 11, that the contribution of the main SubPc absorption peak is clearly visible for all samples. However, in the small wavelengths range, the EQE signal decreases when ZnS/Ag/TiO 2 is the cathode, while a contribution to the current is clearly visible with other TCEs. More precisely, relatively to the maximum of the SubPc contribution, the MoO 3 /Ag/MoO 3 -TCE permits the best answer. These results can be discussed in the light of the spectra of transmission of TCEs and absorption of SubPc and C 60 . We have seen (Figure 9) that there is a decrease of the light transmission of the ZnS/Ag/TiO 2 -TCEs just in the range where the C 60 absorption is maximum, while that of MoO 3 /Ag/MoO 3 is higher than that of ITO, which justifies the shape of the EQE spectra.

Conclusion

In the present work we demonstrate that, in the case of PHJ-IOPVs, the contribution of the C 60 layer to the Jsc, if not principal, is far to be negligible. As a matter of fact, the contribution of SubPc to Jsc is dominant in the visible range, while that of C 60 prevails in the short wavelengths spectral range. These results are corroborated by the optical simulation, which demonstrates the non-negligible contribution of the C 60 acceptor molecule to the absorption of inverse OPVs in the UV-part of the spectral range and therefore to Jsc. Moreover, we show that the transmission spectrum of the TCE is determinant for the shape of the EQE spectrum and thus for Jsc and the OPV efficiency. The alternative ZnS/Ag/TiO 2 indium-free TCE is also proved to be a credible alternative to ITO. However, if TiO 2 is a good electron collector, due to its weak extraction work, the decrease in its transmission below the wavelength of 400 nm penalizes the PHJ-OPVs performances.

Figure 1 :

 1 Figure 1: Device energy diagram

  and CODEX. The different layers evocated above deposited such as, in the classical configuration, we had Glass/TCE/MoO 3 and in the inverted geometry Glass/Alq 3 /C 60 /SubPc/MoO 3 /Al (supporting information s for both OPV configurations are shown in Figure evice energy diagrams for both OPV configurations: (a) COPV, (b) IOPV. OPVs were characterized by J-V, External Quantum Efficiency (EQE) measurements (these experimental techniques are presented in the with the thicknesses 31 35 nm/11 nm/15 nm

3 /

 3 SubPc/C 60 /Alq 3 /Al supporting information S2). s for both OPV configurations are shown in Figure 1. s for both OPV configurations: (a) COPV, (b) IOPV. uantum Efficiency (EQE) techniques are presented in the 3. Experimental results and discussion 3.1 Optimization of the performances of the inverted OPVs, ITO being the transparent electrode.

Figure 2 :

 2 Figure 2: Optimization of the J-V characteristics of inverted OPV by purifying SubPc through deposition cycles: 1 cycle (▼), 3 cycles (▲), 8 cycles (•) and 10 cycles (■) (the filled symbols correspond to dark conditions while the open symbols correspond to light AM 1.5 conditions).

3. 2

 2 Classical and inverted OPVs using ITO as transparent electrode

Figure 3 :

 3 Figure 3: J-V characteristics of IOPV (■) and COPV (▼) (the filled symbols correspond to dark conditions while the open symbols correspond to light AM 1.5 conditions).

Figure 4 :

 4 Figure 4: Typical EQE spectra of IOPV (■) and COPV (•).

3 Figure 5 :

 35 Figure 5: Optical absorption spectra of SubPc ( •• -•• -•• ), C 60 (••••••), ITO/Alq 3 /C 60 /SubPc/MoO 3 (ꟷꟷꟷ) and ITO/MoO 3 /SubPc/C 60 /Alq 3 (--).

Figure 6 :Figure 7 :

 67 Figure 6: Simulated (dashed line) and measured (full line) Transmittance (T), for Glass/ZnS(31 nm)/Ag(10 nm)/TiOx(31 nm) electrode.

Figure 8 :

 8 Figure 8: Absorption of the only (SubPc+C 60 ) active materials inside inverted (IOPV) and classical (COPV) design.

  of C 60 is situated in the very short wavelengths range (below 400 nm), with the second absorption peak of SubPc.

Figure 9 :

 9 Figure 9: Typical transmission spectra of MoO 3 /Ag/MoO 3 (--), ZnS/Ag/TiO 2 (ꟷ ꟷꟷ) and ITO (ꟷ • • ꟷ) and absorption spectra of C 60 (ꟷ • ꟷ) and SubPc (••••).

Figure 10 :

 10 Figure 10: J-V characteristics of IOPV with ZnS/Ag/TiO 2 /Alq 3 (■), MoO 3 /Ag/MoO 3 /Alq 3 (•) and MoO 3 /Ag/MoO 3 /Ca/Alq 3 (▼) as cathode/ETL (the filled symbols correspond to dark conditions while the open symbols correspond to light AM 1.5 conditions).

MoO 3 /

 3 Ag/MoO 3 /Ca/Alq 3 MoO 3 /Ag/MoO 3 /Alq 3

Figure 11 :

 11 Figure 11: Typical EQE spectra of IOPV with different cathodes: (■)ZnS/Ag/TiO 2 ,(•) ITO and (▲)MoO 3 /Ag/MoO 3 .

Table 1 :

 1

	Cathode	ETL	Voc (V)	Jsc	FF (%)	 (%)
				(mA/cm2)
	ZnS/Ag/TiO 2	Alq 3	0.80	4.00	36	1.15
	MoO 3 /Ag/MoO 3	Alq 3	0.40	0.32	31	0.04
	MoO 3 /Ag/MoO 3 Ca/Alq 3	0.86	2.71	48	1.10

Typical parameters of IOPVs (Cathode/ETL/C 60 /SubPc/MoO 3 /Al) with different cathodes.

Acknowledgements

The authors acknowledge funding from the European Community ERANETMED_ENERG-11-196: Project NInFFE "New Indium Free Flexible Electrode".