Andrea Paluselli

Yann Aminot

François Galgani

Sopheak Net

Richard Sempéré
email: richard.sempere@mio.osupytheas.fr

Occurrence of phthalate acid esters (PAEs) in the northwestern Mediterranean Sea and the Rhone River

Keywords:

Phthalate acid esters (PAEs) which are mainly anthropogenic molecules with endocrine disrupting effects in animals and humans, have been detected in terrestrial and aquatic environments. However, little is known about their distribution in the Mediterranean Sea, mainly because of analytical difficulties and the high possibility of ambient sample contamination. Here, we report the optimization of an existing protocol for the determination of PAEs in seawater and freshwater samples, as well as the first estimation of the source and distribution of phthalates acid esters (PAEs) in coastal waters from the NW Mediterranean Sea. By passing 1 L of sample through glass cartridges packed with 200 mg of Oasis HLB and eluted with 6 mL of ethyl acetate followed by gas chromatography and mass spectrometry (GC/ MS) analyses, the recoveries for DMP, DEP, DPP, DiBP, DnBP, BzBP, DEHP and DnOP were 101, 98, 115, 110, 99, 98, 103 and 95%, respectively, with acceptable blank values (below 0.4-4.0% of the masses measured in different seawater samples). By using this method, we detected PAEs in the Marseilles coastal area, offshore (2000 m depth) and in the Rhone River with total concentrations ranging from 75.3 ng/L offshore in surface water to 1207.1 ng/L a few meters above the bottom of the Marseilles Bay. High concentrations were also observed in deep waters offshore (310.2 ng/L) as well as in the Rhone River (615.1 ng/L). These results suggest that Marseilles urban area, Rhone River and sediment are potential sources of PAEs in the areas studied.

In the Rhone River, DEHP was the most abundant PAE (66.1%) followed by DiBP (20.5%) and DnBP (6.6%), whereas a gradual change was observed in the plume of the river with increasing salinity. In the Marseilles Bay, DiBP was the most abundant PAE at the surface (47.3% of total PAEs) followed by DEHP (22.7%) and DnBP (19.1%), whereas DnBP was predominant (38.6%) a few meters above the bottom. By contrast, DEHP was the dominant species in the first 50 m (66.9-76.7%) offshore followed by DnBP (10.9-15.2%), whereas DnBP was the most abundant (57.0-72.6%) followed by DEHP (20.1-33.1%) in the deepest waters. This study suggests that in addition to direct PAEs injection in marine waters, different processes regulate PAE distribution in Mediterranean Sea including photochemical oxidation reactions, bacterial degradation and possible diffusion following release from marine litter near the bottom.

Introduction

Phthalates or phthalic acid ester (PAEs) with low volatility and low water solubility are colorless and odorless. They are mainly produced by anthropogenic activity such as the manufacture and processing of plastic products [START_REF] Simoneit | Combustion products of plastics as indicators for refuse burning in the atmosphere[END_REF][START_REF] Serôdio | Considerations on ultra-trace analysis of phthalates in drinking water[END_REF][START_REF] Lyche | Reproductive and developmental toxicity of phthalates[END_REF]. Their annual production reached 8 million tons in 2011 (Net et al., 2015a(Net et al., , 2015b and references therein) and they represent approximately 70% of the world consumption of plasticizer in 2014 [START_REF] Ceh | Acrylic Resins and Plastics[END_REF]. The structure of PAEs consists of a di-alkyl ester of o-benzenedicarboxylic acid. The most commercially significant PAEs, sorted by increasing number of carbons in the ester side chains, are dimethyl phthalate (DMP), diethyl phthalate (DEP), di-n-butyl phthalate (DnBP), benzylbutyl phthalate (BzBP), di-ethylhexyl phthalate (DEHP) and din-octyl phthalate (DnOP) the later being the most abundant [START_REF] Holahan | Phthalates and neurotoxic effects on hippocampal network plasticity[END_REF]. PAEs of lower molecular weight are used as essential components of solvents, adhesives, waxes, pharmaceutical products, insecticide materials and cosmetics, whereas PAEs of higher molecular weight are instead used to improve the flexibility and handling of industrial materials [START_REF] Iarc | Monograph on the evaluation of carcinogenic risk to humans: some industrial chemicals, di(2-ethylhexyl) phthalate[END_REF][START_REF] Halden | Plastics and health risks[END_REF]. PAEs are mixed with the polymers and are not covalently bound to them. During aging, they are likely to migrate out of the plastic and to be released directly into the environment or inside an animal's stomach or tissue [START_REF] Andrady | Microplastics in the marine environment[END_REF].

As widely used compounds, PAEs are ubiquitous in terrestrial, marine and atmospheric environments [START_REF] Xie | Occurrence and air-sea exchange of phthalates in the Arctic[END_REF][START_REF] Fu | Organic molecular composition of marine aerosols over the Arctic Ocean in summer: contributions of primary emission and secondary aerosol formation[END_REF][START_REF] Dargnat | Phthalate removal throughout wastewater treatment plant: case study of Marne Aval station (France)[END_REF][START_REF] Gao | Occurrence and fate of phthalate esters in fullscale domestic wastewater treatment plants and their impact on receiving waters along the Songhua River in China[END_REF][START_REF] Staples | The environmental fate of phthalate esters: a literature review[END_REF][START_REF] Net | Case study of PAHs, Me-PAHs, PCBs, phthalates and pesticides contamination in the Somme River Water, France[END_REF]Net et al., 2015aNet et al., , 2015b;;[START_REF] Wang | Early life phthalate exposure and atopic disorders in children: a prospective birth cohort study[END_REF]. PAEs have been detected in surface seawater [START_REF] Turner | The behaviour of di-(2-ethylhexyl) phthalate in estuaries[END_REF][START_REF] Horn | Plasticizer metabolites in the environment[END_REF][START_REF] Brossa | Occurrence of twenty-six endocrine-disrupting compounds in environmental water samples from Catalogna, Spain[END_REF], in indoor air [START_REF] Becker | DEHP metabolites in urine of children and DEHP in house dust[END_REF], in urban environments in soils [START_REF] Ferreira | Ex-situ bioremediation of Brazilian soil contaminated with plasticizers process wastes[END_REF], in urban lakes [START_REF] Zeng | Occurrence of phthalate esters in water and sediment of urban lakes in a subtropical city, Guangzhou, South China[END_REF], in drinking water [START_REF] Yang | Monitoring and removal of residual phthalate esters and pharmaceuticals in the drinking water of Kaohsiung City, Taiwan[END_REF], in terrestrial and marine remote aerosols [START_REF] Xie | Development and validation of a method for the determination of trace alkylphenols and phthalates in the atmosphere[END_REF][START_REF] Xie | Occurrence and air-sea exchange of phthalates in the Arctic[END_REF][START_REF] Fu | Organic molecular compositions and temporal variations of summertime mountain aerosols over Mt. Tai, North China Plain[END_REF][START_REF] Fu | Photochemical and other source of organic compounds in the Canadian high Arctic aerosol pollution during winter-spring[END_REF][START_REF] Fu | Organic molecular composition of marine aerosols over the Arctic Ocean in summer: contributions of primary emission and secondary aerosol formation[END_REF], and in marine sediment, as well as in biota, including several aquatic species [START_REF] Wofford | Bioaccumulation and metabolism of phthalate esters by oysters, brown shrimp, and sheepshead minnows[END_REF], algae [START_REF] Gao | Biodegradation of phthalate acid esters by different marine microalgal species[END_REF], and the tissues of mammals [START_REF] Staples | The environmental fate of phthalate esters: a literature review[END_REF] and humans [START_REF] Liou | Monitoring of PAEMs and betaagonists in urine for a small group of experimental subjects and PAEs and betaagonists in drinking water consumed by the same subjects[END_REF]. PAEs can be degraded by some microalgal species [START_REF] Gao | Biodegradation of phthalate acid esters by different marine microalgal species[END_REF], fish and shellfish [START_REF] Munshi | Toxicity of phthalate esters in fish and shellfish from Virginia Beach using matrix solid phase dispersion (MSPD) and GC-MS[END_REF], bacteria [START_REF] Chang | Biodegradation of phthalates ester by two bacteria strains[END_REF] and intracellular and extracellular algal enzymes [START_REF] Gao | Biodegradation of phthalate acid esters by different marine microalgal species[END_REF] as well as by the combined action of UV and H 2 O 2 [START_REF] Xu | Photochemical degradation of diethyl phthalate with UV/H 2 O 2[END_REF]. PAEs are of particular concern to humans since they are considered as emerging environmental contaminants due to the endocrine disruption and carcinogenic effects detected in laboratory animal experiments and in humans [START_REF] Crisp | Environmental endocrine disruption: an effects assessment and analysis[END_REF][START_REF] Latini | Monitoring phthalate exposure in humans[END_REF][START_REF] Kamrin | Phthalate risks, phthalate regulation, and public health: a review[END_REF][START_REF] Meeker | Phthalates and other additives in plastics: human exposure and associated health outcomes[END_REF].

The Mediterranean Sea is a semi-enclosed basin covering an area of 2.5 million km 2 with a slow turnover time of $80 years implying a sensitive response to anthropogenic impact (The [START_REF] Mermex | Marine ecosystems responses to climatic and anthropogenic forcings in the Mediterranean[END_REF]. Urbanization has been growing particularly along the coastline, with the result of a substantial modification of the coast and adverse effects on the quality of the environment [START_REF] Pergent | Mediterranean Seagrass Meadows: Resilience and Contribution to Climate Change Mitigation[END_REF]. There is a large range of industrial activities spread all along the Mediterranean basin, and a number of highly industrialized spots that are concentrated mainly in the northwestern Mediterranean Sea [START_REF] Zorita | Assessment of biological effects of environmental pollution along the NW Mediterranean Sea using mussels as sentinel organisms[END_REF]. All these activities constitute a source of pollution through direct disposal, continental runoff and atmospheric transport (UNEP [START_REF] Chemicals | Regionally Based Assessment of Persistent Toxic Substances[END_REF]The Mermex group, 2011 and references therein). These industrial, agricultural and domestic activities generate a large volume of wastewater, which provides nutrients, trace metals [START_REF] Cossa | Chronology of anthropogenic trace elements accumulation in the Rhone pro-delta sediments (northwestern Mediterranean) during the last 400 years[END_REF] organic matter, persistent organic pollutants (POPs) and related anthropogenic contaminants such as polycyclic aromatic hydrocarbons (PAHs) and polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs) (The [START_REF] Mermex | Marine ecosystems responses to climatic and anthropogenic forcings in the Mediterranean[END_REF][START_REF] Berrojalbiz | Biogeochemical and physical controls on concentrations of polycyclic aromatic hydrocarbons in water and plankton of the Mediterranean and Black Sea[END_REF][START_REF] Guigue | Occurrence and distribution of hydrocarbons in the surface microlayer and subsurface water from the urban coastal marine area off Marseilles, Northwestern Mediterranean Sea[END_REF][START_REF] Guigue | Spatial and seasonal variabilities of dissolved hydrocarbons in surface waters from the Northwestern Mediterranean Sea: results from one year intensive sampling[END_REF][START_REF] Castro-Jimenez | PCDD/F and PCB multi-media ambient concentrations, congener patterns and occurrence in a Mediterranean coastal lagoon (Etang de Thau, France)[END_REF][START_REF] Halden | Plastics and health risks[END_REF], 2012) that impact the coastal marine ecosystems. The Mediterranean Sea is also impacted by marine litter and microplastics [START_REF] Deudero | Mediterranean marine biodiversity under threat: reviewing influence of marine litter on species[END_REF][START_REF] Cozar | Plastic accumulation in the Mediterranean Sea[END_REF][START_REF] Faure | Plastic pollution in Swiss surface waters: nature and concentrations, interaction with pollutants[END_REF][START_REF] Pedrotti | Changes in the floating plastic pollution of the Mediterranean Sea in relation to the distance to land[END_REF][START_REF] Suaria | The Mediterranean Plastic Soup: synthetic polymers in Mediterranean surface waters[END_REF][START_REF] Schmidt | Occurrence of microplastics in surface waters of the Gulf of Lion (NW Mediterranean Sea)[END_REF], which can supply PAEs to surrounding waters during their degradation [START_REF] Andrady | Microplastics in the marine environment[END_REF]. Although PAEs have already been detected in aquatic environments [START_REF] Xie | Atmospheric concentrations and air-sea exchanges of phthalates in the North Sea (German Bight)[END_REF]Net et al., 2015aNet et al., , 2015b)), only a few studies have reported the occurrence of individual phthalates in the Mediterranean Sea [START_REF] Sanchez-Avila | Organic micropollutants in coastal waters from NW Mediterranean Sea: sources distribution and potential risk[END_REF][START_REF] Brossa | Occurrence of twenty-six endocrine-disrupting compounds in environmental water samples from Catalogna, Spain[END_REF], and there is no study dealing with the distribution of a complete series of PAEs in the Mediterranean basin, including the sea and rivers, mainly because of analytical difficulties.

Indeed, PAE analysis of environmental samples is a challenging task because of the low concentrations, which is associated with the risk of contamination during sampling and processing in relation with their ubiquity in the laboratory atmosphere, tools, glassware and various containers (Net et al., 2015b and references therein). For environmental waters, several extraction procedures have already been used, including liquid-liquid (EPA, 1996;[START_REF] Zeng | Occurrence of phthalate esters in water and sediment of urban lakes in a subtropical city, Guangzhou, South China[END_REF][START_REF] Li | Rapid analysis of phthalates in beverage and alcoholic samples by multi-walled carbon nanotubes/silica reinforced hollow fibre-solid phase microextraction[END_REF][START_REF] Gao | Occurrence and fate of phthalate esters in fullscale domestic wastewater treatment plants and their impact on receiving waters along the Songhua River in China[END_REF], solid-liquid (SLE), solid phase (SPE), magnetic (MSPE) dispersive (DSPE), solid phase micro extraction (SPME) [START_REF] Zhang | Diethylhexyl phthalate exposure impairs follicular development and affects oocyte maturation in the mouse[END_REF][START_REF] Liou | Monitoring of PAEMs and betaagonists in urine for a small group of experimental subjects and PAEs and betaagonists in drinking water consumed by the same subjects[END_REF][START_REF] Yang | Monitoring and removal of residual phthalate esters and pharmaceuticals in the drinking water of Kaohsiung City, Taiwan[END_REF] and high volume sample techniques with PAD-2 resin columns and Soxhlet extractor [START_REF] Xie | Atmospheric concentrations and air-sea exchanges of phthalates in the North Sea (German Bight)[END_REF]. SPE appears to be an appropriate technique for seawater because of its ease of implementation, its ability to save time and solvent (compared to liquidliquid extraction) and the decreased contamination [START_REF] Zhang | Diethylhexyl phthalate exposure impairs follicular development and affects oocyte maturation in the mouse[END_REF][START_REF] Liou | Monitoring of PAEMs and betaagonists in urine for a small group of experimental subjects and PAEs and betaagonists in drinking water consumed by the same subjects[END_REF][START_REF] Yang | Monitoring and removal of residual phthalate esters and pharmaceuticals in the drinking water of Kaohsiung City, Taiwan[END_REF]Net et al., 2015b). However, the SPE protocol uses a tube made of polyethylene or polypropylene, which can release DiBP, DnBP and DEHP [START_REF] Tienpont | Determination of Phthalates in Environmental, Food and Biomatrices[END_REF][START_REF] Fasano | Migration of phthalates, alkylphenols, bisphenol A and di(2-ethylhexyl)adipate from food packaging[END_REF][START_REF] Bonini | Extraction and gas chromatographic evaluation of plasticizers content in food packing film[END_REF]Paluselli unpublished results). This technique has been used by a great number of reported protocols, concerning the type, amount and conditioning of the phase [START_REF] Liou | Monitoring of PAEMs and betaagonists in urine for a small group of experimental subjects and PAEs and betaagonists in drinking water consumed by the same subjects[END_REF][START_REF] Sanchez-Avila | Development of a multi-residue method for the determination of organic micropollutants in water, sediment and mussels using gas chromatography-tandem mass spectrometry[END_REF], 2012;[START_REF] Yang | Monitoring and removal of residual phthalate esters and pharmaceuticals in the drinking water of Kaohsiung City, Taiwan[END_REF] and the sample volume as well as the volume and type of solvent [START_REF] He | Spatio-temporal distributions and the ecological and health risks of phthalate esters (PAEs) in the surface water of a large, shallow Chinese lake[END_REF][START_REF] Sanchez-Avila | Development of a multi-residue method for the determination of organic micropollutants in water, sediment and mussels using gas chromatography-tandem mass spectrometry[END_REF], 2012;[START_REF] Yang | Monitoring and removal of residual phthalate esters and pharmaceuticals in the drinking water of Kaohsiung City, Taiwan[END_REF][START_REF] Liou | Monitoring of PAEMs and betaagonists in urine for a small group of experimental subjects and PAEs and betaagonists in drinking water consumed by the same subjects[END_REF]. Previous protocols have shown a recovery variability ranging from 50-60% to 120-150% [START_REF] Del Carlo | Determination of phthalate esters in wine using solid-phase extraction and gas chromatography-mass spectrometry[END_REF][START_REF] Wang | Distribution of phthalic acid esters in Wuhan section of the Yangtze River, China[END_REF][START_REF] He | Spatio-temporal distributions and the ecological and health risks of phthalate esters (PAEs) in the surface water of a large, shallow Chinese lake[END_REF][START_REF] Sanchez-Avila | Development of a multi-residue method for the determination of organic micropollutants in water, sediment and mussels using gas chromatography-tandem mass spectrometry[END_REF], especially for DEHP, which ranges from 60-80% [START_REF] He | Spatio-temporal distributions and the ecological and health risks of phthalate esters (PAEs) in the surface water of a large, shallow Chinese lake[END_REF] to 151% [START_REF] Sanchez-Avila | Development of a multi-residue method for the determination of organic micropollutants in water, sediment and mussels using gas chromatography-tandem mass spectrometry[END_REF].

The objectives of this study are to examine, for the first time, the concentration levels and distribution of a homologous series of PAEs in the coastal northwestern Mediterranean Sea. This environmental study was associated with an optimization of a detection protocol for PAEs at trace level in seawater. Based on previous studies [START_REF] Liou | Monitoring of PAEMs and betaagonists in urine for a small group of experimental subjects and PAEs and betaagonists in drinking water consumed by the same subjects[END_REF][START_REF] Gao | Biodegradation of phthalate acid esters by different marine microalgal species[END_REF][START_REF] Yang | Monitoring and removal of residual phthalate esters and pharmaceuticals in the drinking water of Kaohsiung City, Taiwan[END_REF], conditions for the SPE extraction of PAEs have been modified and optimized for seawater. The method we report here was tested in seawater spiked with authentic standards and then applied to coastal and offshore Mediterranean seawater as well as to Rhone River freshwater in the framework of the 'Particule-MERMEX' project.

Experimental section

Materials and reagents

The selected phthalates are dimethyl phthalate (DMP), diethyl phthalate (DEP), dipropyl phthalate (DPP), di-isobutyl phthalate (DiBP), di-n-butyl phthalate (DnBP), benzylbutyl phthalate (BzBP), di-ethylhexyl phthalate (DEHP) and di-n-octyl phthalate (DnOP). The deuterated internal standards were DEP-d4, DnBP-d4 and DEHP-d4. All native and labeled standards were of high purity grade (>98%) and were purchased from Supelco (Bellefonte, USA). Stock solutions were prepared in hexane and stored in the dark at 4 °C. Working solutions (unlabeled and labeled standards) were prepared by dilution of these solutions to 20 mg/L. All solvents were glass-distilled grade and supplied by Rathburn Chemicals Ltd. (Walkerburn, UK). Ultra-pure water was produced on-site by a Milli-Q system, Millipore (Molsheim, France) with a specific resistivity of 18.2 MXÁcm (25 °C) and a total organic carbon content of <2 ppb. The extractions were carried out in controlled air conditions in an ISO class 6 cleanroom (temperature: 22 °C; SAS pressure: +15 Pa; SAS brewing rate: 30 vol/h; lab pressure: +30 Pa; brewing rate: 50 vol/h). All material was produced in borosilicate glass and that was previously kept in an acid bath overnight (10% hydrochloric acid), rinsed with ultrapure water, combusted at 450 °C for 6 h and rinsed with methanol and dichloromethane just before use.

Solid phase extraction (SPE)

The protocol was adapted from that of [START_REF] Yang | Monitoring and removal of residual phthalate esters and pharmaceuticals in the drinking water of Kaohsiung City, Taiwan[END_REF] and [START_REF] Liou | Monitoring of PAEMs and betaagonists in urine for a small group of experimental subjects and PAEs and betaagonists in drinking water consumed by the same subjects[END_REF] and optimized for seawater in this study. The mentioned methods were modified in order to perform easy and rapid analysis on a large-scale with low contamination and to improve the recovery. The amount of Oasis HLB phase, the elution solvent type and volume, the sample volume, deuterated internal standards and glass materials were considered to improve the recoveries. Extraction cartridges were prepared in our laboratory with 200 mg of Oasis HLB sorbent (Waters Corporation, 30 mm, 100 g) in a precombusted 6 mL glass reaction tube equipped with methanol-precleaned Teflon frits (Supelco, Sigma-Aldrich). SPE was conducted on a Visiprep vacuum manifold from Sigma-Aldrich (Saint Quentin Fallavier, France). Before sample extraction, the SPE cartridges were sequentially cleaned three times with 5 mL of acetone and 5 mL of dichloromethane. The cartridges were then conditioned with 5 mL of ethyl acetate, 5 mL of acetone and 5 mL of ultrapure water prior to sample processing. Seawater samples were spiked with DEP-d4, DnBP-d4 and DEHP-d4 at 50, 200 and 400 ng/L, respectively, and percolated onto the cartridges under vacuum at a flow rate of 1-2 drops per second. After percolation, the cartridge was washed with 5 mL of ultrapure water to remove the remaining salt and dried under vacuum for 1 h. The retained phthalates were then eluted with 2 Â 3 mL of ethyl acetate into precombusted (450 °C, 6 h) 10 mL vials. Ethyl acetate was then gently evaporated under a gentle stream of nitrogen (purity >99,995%) at room temperature to a final volume of 200 mL. The final samples were immediately closed with Teflon-lined screw caps, wrapped with Teflon ribbon and stored in the dark at À20 °C until analysis.

GC-MS analysis

Analysis was performed using an Agilent Technologies 6850 GC system coupled to an Agilent Technologies 5975C mass spectrometer (GC/MS) operated with electron impact ionization (70 eV). Chromatographic separation was achieved on an Agilent HP-5MS capillary column (30 m  0.25 mm, 0.25 mm film thickness) with the GC oven programmed at 70 °C for 1 min and then up to 230 °C at 30 °C/min, to 250 °C at 8 °C/min and to 280 °C at 30 °C/ min with a final isothermal hold for 7.50 min. Helium was used as the carrier gas at a flow rate of 1.1 mL/min. The sample was manually injected (2 mL) in splitless mode at an injector temperature of 250 °C. A Merlin Microseal injector was used as a low carryover septum and a gas purifier (Charcoal, CP17972) was fitted on the carrier gas line to prevent contamination during injection. Data were acquired in selected ion monitoring (SIM) mode for increased sensitivity (SIM parameters given in Table 1) with a dwell time of 100 ms. Two fragment ions were monitored per compound, with the most abundant ion used for quantification, except for DMP, DPP, DiBP and DnBP, which were ionized in a unique fragment. Data were collected and analyzed with the Agilent ChemStation software. The PAEs were quantified by isotopic dilution of 3 deuterated PAEs. The use of internal standard homologues to the target molecules has been shown to improve the recoveries by accurately compensating for loss of analyte during sample treatment [START_REF] Aminot | Development and application of a multi-residue method for the determination of 53 pharmaceuticals in water, sediment, and suspended solids using liquid chromatography-tandem mass spectrometry[END_REF]. The response factors were assessed for each analytical sequence from freshly prepared standards.

Application to seawater and freshwater samples

Seawater samples were collected from SUNMEX station in the northwestern Mediterranean Sea, Bay of Marseilles (43°18 0 N, 05°22 0 E) in October 2014 and offshore at 'Antares station' (42°47 0 856N; 6°04 0 395E) in March 2015 (Fig. 1), with a 12 L GO-Flo bottle previously rinsed with 1% hydrochloric acid and ultrapure water to avoid contamination. The Go-Flo bottle tested with Milli-Q water showed only trace levels of PAEs which were not substracted from our dissolved PAEs concentrations in this study. Immediately after collection, seawater samples were transferred into 5 L precombusted glass bottles. Additional samples were directly collected in the Rhone River and Rhone River plume in March 2014 using 5 L precombusted glass bottles. General information on the Marseilles Bay, the Antares station and the Rhone River were given by [START_REF] Sempéré | Variability of solar radiation and CDOM in surface coastal waters of the northwestern Mediterranean Sea[END_REF], [START_REF] Martini | Bacteria as part of bioluminescence emission at the deep ANTARES station (North-Western Mediterranean Sea) during a one-year survey[END_REF] and [START_REF] Sempéré | Carbon inputs of the Rhône River to the Mediterranean Sea: biogeochemical implications[END_REF], respectively. Less than 2 h after collection, the samples were filtered through precombusted (450 °C for 6 h) GFF filters (47 mm filter diameter, rinsed with 2 L of Milli-Q water and 150 mL of sample prior to filtration) in a precombusted glass apparatus directly in a filtered-air clean laboratory. Milli-Q water rinse limit additional contamination [START_REF] Yoro | Dissolved organic carbon contamination induced by filters and storage bottles[END_REF]. After filtration, duplicate subsamples (10 mL) were poisoned with sulfuric acid to approximately pH 2 to avoid any biological activity and stored in the dark in precombusted Teflon-capped glass vials for dissolved organic carbon (DOC) analysis. DOC concentrations were measured using a Shimadzu TOC-5000 carbon analyzer [START_REF] Sempéré | Distribution and bacterial availability of dissolved neutral sugars in the South East Pacific[END_REF]. Some seawater samples (Bay of Marseilles) were spiked with authentic PAEs (from 5 to 800 ng/L final concentration) to perform recovery experiments. All the seawater analyses were conducted in duplicate in the laboratory in Marseilles.

Results and discussion

Extraction of PAEs in seawater samples (optimization of the method)

Numerous factors are known to significantly influence the efficiency of SPE extraction of phthalates from water. The most influential factors are elution the solvent, sample volume and sorbent mass. To increase the recovery, the elution solvent, sample volume and sorbent mass were studied in order to optimize efficient and reliable conditions for SPE extraction of seawater samples.

Elution solvent effect

First, 10 mL of 20 mg/L PAE solution was passed through the cartridges, and two fractions of elution were collected. The first fraction was eluted with 5 mL of either acetone, dichloromethane, ethyl acetate, methanol or n-hexane, and the second fraction was systematically eluted with 5 mL of dichloromethane in order to elute the remaining PAEs off the cartridge. Internal standards were spiked directly in the eluted fraction. The repartition of PAEs between the two fractions is given in Fig. 2. Acetone, dichloromethane, ethyl acetate and methanol yielded acceptable recoveries ranging from 55 to 100% while n-hexane was satisfactory only for DEHP and DnOP (79-85%), with 12% for DMP, 36% for DEP and 1% for BzBP. n-Hexane is a better solvent for aliphatic compounds, unlike DMP, DEP (short carbon chain) and BzBP (higher aromaticity).

Acetone and methanol showed a similar elution strength with recoveries of 80-100% for DMP, DEP, DPP, DiBP, DnBP and BzBP in the first fraction and 70% and 50% for DEHP and DnOP, respectively. This result is consistent with the studies of [START_REF] Liou | Monitoring of PAEMs and betaagonists in urine for a small group of experimental subjects and PAEs and betaagonists in drinking water consumed by the same subjects[END_REF] and [START_REF] Yang | Monitoring and removal of residual phthalate esters and pharmaceuticals in the drinking water of Kaohsiung City, Taiwan[END_REF], which performed the elution with 6 mL of methanol. Dichloromethane and ethyl acetate both indicated high recoveries for all PAEs in the first fraction and were the most appropriate elution solvents. Dichloromethane was the most appropriate elution solvent for DMP (100%), DEP (98%) and DPP (97%), whereas ethyl acetate was the most appropriate for the elution of DiBP (91%), DnBP (92%), BzBP (85%) and DEHP (90%). DnOP was efficiently eluted (100%) with both solvents. Previously, published protocols used 10 mL of dichloromethane/nhexane [START_REF] He | Spatio-temporal distributions and the ecological and health risks of phthalate esters (PAEs) in the surface water of a large, shallow Chinese lake[END_REF] or 10 mL of acetone/dichloromethane [START_REF] Sanchez-Avila | Development of a multi-residue method for the determination of organic micropollutants in water, sediment and mussels using gas chromatography-tandem mass spectrometry[END_REF], 2012) as the elution solvent. When comparing their recoveries and those from our developed protocol, the combination of dichloromethane with acetone seems to be more effective for PAE elution than its combination with n-hexane. In these protocols, the ability to elute PAEs from the cartridges was improved by increasing the elution solvent volume to 10 and 20 mL, respectively. According to our results, ethyl acetate is the best solvent for PAE elution.

Seawater volume and sorbent mass effects

As high SPE enrichment factors are required to compensate for the likely low PAE levels in seawater, it is important to ensure that the analytes do not breakthrough the sorbent at high percolation volumes. Increasing volumes (100 mL, 200 mL, 500 mL and 1000 mL) of seawater collected at Sunmex station at a 1.5 m depth were spiked with 10 mL of 20 mg/L PAEs and percolated through either 200 mg [START_REF] Liou | Monitoring of PAEMs and betaagonists in urine for a small group of experimental subjects and PAEs and betaagonists in drinking water consumed by the same subjects[END_REF] or 400 mg of sorbent. For each experimental condition, two additional samples were kept unspiked and were processed in the same conditions (blank samples). The recoveries were calculated as the difference between the concentration in the spiked and unspiked samples. In this experiment, the deuterated internal standards DEP-d4, DnBP-d4 and DEHP-d4 were spiked after the elution of the solvent. The recoveries of the extractions obtained with 4 different volumes of spiked seawater for 200 and 400 mg of HLB sorbent are given in Table 2.

The volume of the sample does not significantly affect the recovery of most of the PAEs, indicating the absence of breakthrough. Only DEP and DnBP were marginally affected by increasing sample volumes. The recovery of DEP increased from 60% with 100 mL to 93% with all other volumes on 200 mg of sorbent. However, the recovery was stable between 93 and 100% on 400 mg. DnBP showed the opposite trend, where the recovery decreased from 75-79% with 100-200 mL to 63-66% with 500-1000 mL on both 200 mg and 400 mg of sorbent. Acceptable recoveries (93-120%) were determined for DMP, DPP, DiBP and BzBP under all experimental conditions. The two most hydrophobic PAEs DEHP and DnOP exhibited lower recoveries regardless of the experimental conditions, attributed to their potential for adsorption onto glassware. Such losses were corrected for by the use of deuterated DEHP as an internal standard (Table 3).

Above all, the volume of the sample did not significantly influence the performance of the HLB sorbent toward PAEs. Similar results were observed for the sorbent mass. The DnOP concentration in seawater is expected to be low, indicating the need to increase the enrichment factor by extracting 1 L of sample. BzBP is also characterized by low concentrations in seawater and by relatively high instrumental detection limits, confirming the need for a large volume of sample. As the differences between 200 and 400 mg of sorbent did not justify the use of double the amount of sorbent, 200 mg of sorbent was selected to reduce the price and time for a single extraction. Previously, published protocols reported the use of 1 L of matrix, but different amounts of solid phase, from 200 mg [START_REF] Liou | Monitoring of PAEMs and betaagonists in urine for a small group of experimental subjects and PAEs and betaagonists in drinking water consumed by the same subjects[END_REF] to 500 mg [START_REF] Yang | Monitoring and removal of residual phthalate esters and pharmaceuticals in the drinking water of Kaohsiung City, Taiwan[END_REF]. According to the results, our optimized conditions were 200 mg of HLB phase prepared in a combusted glass tube, 1 L of sample volume, and 6 mL of ethyl acetate as the elution solvent, split in two steps.

Method validation

Blank, response linearity and extraction validation

Procedural blank samples were examined by passing dry air and ethyl acetate directly through the precleaned cartridge and measuring the PAEs directly in the elution solution. Linearity was investigated over an eight-point calibration ranging from 50 to 6000 pg injected. Instrumental repeatability was measured through the relative standard deviation of 10 injections of the same standard solution, corresponding to 100 pg injected, whereas instrumental detection limits (IDLs) were extrapolated for a signal-to-noise ratio of 3. The noise region was equivalent to the width of the peak before the peak, and the noise value was taken as the average height (''RMS" mode). The recoveries were calculated from the spiked seawater samples (10 mL of 20 mg/L of PAEs in 1 L) after subtraction of the concentrations measured for the unspiked seawater samples (n = 2). Such experiments were repeated with half the spiked concentration (5 mL of 20 mg/L of PAEs) in seawater samples (n = 2) to study the influence of the concentration level.

The method was applied to 1 L seawater samples to minimize contamination and to allow the detection of low concentrations of BzBP and DnOP in seawater. The samples were passed through 200 mg of HLB sorbent, and the cartridges were eluted twice with 3 mL of ethyl acetate. A typical gas chromatogram of the PAEs isolated from the seawater samples collected in the Marseilles Bay is given in Fig. 3. Identification of the PAEs was confirmed by GC/MS. Despite the precautions taken to prevent contamination, DEP, DiBP and occasionally DnBP were detected in the blanks at levels that remained below 0.4-2%, 2-3% and 0-4% of the masses measured in different seawater samples, respectively. The average recoveries of PAEs spiked in the seawater samples (Table 3) showed accept-able recoveries for all the phthalates, ranging from 97% of DEP to 110% of DiBP. The high recoveries obtained at low spiking levels also validated the appropriate performance of the protocol. To assess the influence of the concentration level on the recoveries, 1 L aliquots of seawater collected in the Bay of Marseilles were spiked at increasing concentration levels (standard addition technique). Different ranges of spiked concentrations were used depending on the initial concentration of the compound in seawater (from 5 to 40 ng/L for DMP, DEP, DPP, BzBP and DnOP and from 40 to 800 ng/L for DiBP, DiBP and DEHP). Excellent correlation coefficients (R 2 > 0.96) and recoveries (between 95 and 116%) indicate that the method is linear and accurate for all compounds, including DEHP (R 2 = 0.987; n = 7) and DnOP (R 2 = 0.962; n = 5), in a range of environmentally realistic concentrations.

Concentration of PAEs in the northwestern Mediterranean Sea and Rhone River

All PAEs, excepted DPP and DnOP, were detected in the Mediterranean seawater and Rhone River samples (Table 4). We found that offshore (Antares site), the concentration of total PAEs ranged from 75.3 ng/L at the surface to 310.2 ng/L in the deepest water, from 244.5 ng/L to 1207.1 ng/L a few meters above the bottom in the Marseilles Bay and from 182.2 to 212.6 ng/L in the Rhone River plume and close to the Marseilles Cortiou water treatment plant outlet. Although particulate PAEs were not measured in this study, previous study indicated that DnBP, BzBP and DEHP particulate concentration may represent the 2, 29 and 43%, respectively, of the dissolved concentration in coastal seawater (North Sea, [START_REF] Xie | Atmospheric concentrations and air-sea exchanges of phthalates in the North Sea (German Bight)[END_REF]. Very limited data are available for the occurrence and fate of dissolved PAEs in marine water, especially in the Mediterranean Sea. DEHP concentrations detected in the seawater samples in our study were in the range of 50.7-296.5 ng/L. These concentrations were in the same range reported by [START_REF] Sanchez-Avila | Organic micropollutants in coastal waters from NW Mediterranean Sea: sources distribution and potential risk[END_REF] for Catalan coastal water (Spain, NW Mediterranean Sea) and by [START_REF] Turner | The behaviour of di-(2-ethylhexyl) phthalate in estuaries[END_REF] for the south coast of the UK, respectively, with range of 31-617 ng/L and 99-2200 ng/ L. However, our results were 10 times lower than those reported by [START_REF] Brossa | Occurrence of twenty-six endocrine-disrupting compounds in environmental water samples from Catalogna, Spain[END_REF] for the Spanish coast. Lower concentrations have been only reported for North Sea with concentration ranges of 0.5-5.3 ng/L for DEHP and 0.02-4.0 ng/L for DMP and DEP [START_REF] Xie | Atmospheric concentrations and air-sea exchanges of phthalates in the North Sea (German Bight)[END_REF].

In Marseilles Bay, the samples were characterized by high concentrations of PAEs in the surface waters, a slight decrease with the depth, probably because of microbial degradation, and then an increase close to the bottom (Table 4). DiBP was the most abundant at the surface (47.3% of total PAEs) followed by DEHP (22.7%) and DnBP (19.1%), whereas DnBP was predominant (38.6%) followed by DiBP (31.8%) and DEHP (24.6%) a few meters above the bottom. DMP, DEP and BzBP were detected as 0.5, 2.5 and 0.8% of total PAEs, respectively (Table 5; Fig. 4). The high concentration at the bottom suggests the resuspension of PAE-rich sediments. As marine litter has already been reported to be abundant near the bottom (M I O unpublished observations) of the Marseilles Bay, it is likely that PAEs in the surficial sediment were released during plastic debris aging through physico-chemical and biotic processes [START_REF] Bakir | Enhanced desorption of persistent organic pollutants from microplastics under simulated physiological conditions[END_REF]. This seems to be more evident for DnBP, whose relative fraction showed a stronger variation. Different PAEs are usually used by industry during plastic polymer production in order to increase flexibility, transparency, durability and longevity of the different materials produced (Net et al., 2015b, and reference therein). DEHP and DnBP are used as plasticizer for soft PVC, PVA and rubber as well as fixative in paint and especially DnBP is essentially used for its viscosity reducing properties and compatibility with non-PVC mixture (ECHA, 2013). It is likely that the PAEs detected above the bottom depend on the type of material that accumulates in the sediment after sinking. In addition, microplastics could come from larger plastic debris that degrades into smaller pieces and microbeads, a type of microplastic that are added as exfoliants to health and beauty products and easily pass though water filtration systems and end up in the sea, could transfer attached pollutants and additive chemicals [START_REF] Browne | Microplastic -an emerging contaminant of potential concern?[END_REF] in the water column. At Cortiou station, near the Marseilles wastewater treatment plant, the surface water PAE concentration averaged at 212.6 ng/L (sd ± 3.4) (Table 4), detected 5-6 m from the Cortiou water treatment plant outlet, which was quite low during the studied period. DiBP was the most abundant PAE (38.4%), followed by DEHP (26.4%), DEP (15.7%) and DnBP (13.5%), with DMP and BzBP representing minor fractions (3.5 and 2.5%, respectively) (Table 5; Fig. 4). The relative abundance was not very different from what we detected in Marseilles Bay, with the only exception of DEP.

Five PAEs were detected offshore in the 'Antares station' water column (Figs. 5 and6; Table 6). DnBP was characterized by a large range of concentrations (8.2-225.5 ng/L), which reached a maximum in deep water (Fig. 6). With the exception of DnBP, the vertical profiles seem to have similar trends. The surface samples were characterized by high concentrations that slightly decrease down to 200-500 m. Then, the concentration started to increase down to 1500 m, where the concentration was similar to that detected in the surface samples. The DnBP vertical profile was characterized by the same trend in the first meter of the water column, decreasing from 13.4 to 8.2 ng/L, but it was followed by a strongly increasing concentration in the deepest samples (87.3-225.5 ng/L). DEHP was the dominant species in the first 50 m (66.9-76.7% of total PAEs) followed by DnBP (10.9-15.2%). Between 100 and 250 m, the relative abundance of DEHP was slightly lower (56.9-65.4%), whereas that of DnBP was higher (25.8-28.5%) up to the maximum salinity and minimum temperature. Under 500 m, DnBP was the most abundant PAE (57.0-72.6%) followed by DEHP (20.1-33.1%). DiBP, DMP and DEP were detected as minor fractions of the total PAEs detected in offshore waters, representing 7.9, 5.7 and 4.4% of total PAEs at the surface and 3.5, 2.3 and 1.5% in deepest waters, respectively (Fig. 5; Table 6). Note that this offshore station area is influenced by the shallow depth Northern Current (NC) with high seasonal variability [START_REF] Millot | Mesoscale and seasonal variabilities of the circulation in the western Mediterranean[END_REF], which may penetrate into the shelf of the Gulf of Lions (Barrier et al., 2016 and references therein). This water originating eastward from the Italian coasts may carry plastic debris [START_REF] Schmidt | Occurrence of microplastics in surface waters of the Gulf of Lion (NW Mediterranean Sea)[END_REF] and dissolved PAEs. The decrease in concentration in the first meter of the water column could be explained by biodegradation processes. The increase in concentration at 200-500 m suggests the production of PAEs from the aging of plastic debris, as was observed in Marseilles Bay. Indeed, the behavior of DnBP in the deepest water seems to reflect the same production observed close to the bottom of the Marseilles area. These data were the first attempt to investigate the vertical distribution of PAEs in the water column.

Rhone River and river mouth

The total PAE concentrations ranged from 615.1 ng/L to 182.2-220.0 ng/L in the Rhone River plume (Table 4) suggesting that the Rhone is clearly a source of PAEs in the coastal area. Previous studies dealing with individual PAEs reported DMP concentrations in the range of 2.5-5.5 ng/L (Sanchez-Avila et al., 2012) in six rivers in Spain (Muga, Fluvià, Ter, Besòs, Llobregat and Ebro River). DEP was detected in different studies in the Seine River (France) in a range slightly higher than our data, from 46-208 ng/L [START_REF] Teil | Occurrence of phthalates diesters in river of the Paris district (France)[END_REF] to 71-181 ng/L [START_REF] Dargnat | Phthalate removal throughout wastewater treatment plant: case study of Marne Aval station (France)[END_REF], as well as 52-284 ng/L in Spain [START_REF] Sanchez-Avila | Organic micropollutants in coastal waters from NW Mediterranean Sea: sources distribution and potential risk[END_REF]. BzBP was also reported at low concentration in the range of 5-23 ng/L [START_REF] Dargnat | Phthalate removal throughout wastewater treatment plant: case study of Marne Aval station (France)[END_REF][START_REF] Sanchez-Avila | Organic micropollutants in coastal waters from NW Mediterranean Sea: sources distribution and potential risk[END_REF]. [START_REF] Dargnat | Phthalate removal throughout wastewater treatment plant: case study of Marne Aval station (France)[END_REF] and [START_REF] Teil | Occurrence of phthalates diesters in river of the Paris district (France)[END_REF] found similar concentrations of DEHP, 161-314 ng/L and 323-779 ng/L, respectively, in the Seine River. We found that DEHP was the most abundant PAE (66.1%) in the Rhone River followed by DiBP (20.5%) and DnBP (6.6%), whereas DMP, DEP and BzBP represented the smallest fractions (Table 5; Fig. 4). The relative abundance of DEHP decreased from 66.1% of the total PAEs in the river (S = 0) to 48.4, 30.7 and 17.8% in the plume stations (S = 10-28). The opposite trend was observed for all other PAEs, especially for DnBP and DiBP, with DnBP values increasing from 6.6 to 11.9% and from 20.5 to 62.0% for DiBP. A similar trend was also observed for DMP and DEP. Changes in the relative abundance from the river to the plume could be due to different selective bacterial degradation and/or photodegradation processes occurring in the salty water. Only DiBP showed, after the concentration first decreased probably due to dilution effect, an increase in concentration and in relative abundance. This could be caused by some atmospheric input of this compound as it was already supposed in the Bay of Marseilles.

Conclusion

A reliable analytical technique for measuring PAEs in natural seawater samples at trace levels was slightly improved and allowed determination of PAEs in seawater. This technique consists of the adsorption on PAEs from acidified seawater samples onto 200 mg Oasis HLB glass cartridges followed by elution with ethyl acetate and determination by GC/MS. The recoveries of the PAEs from spiked seawater samples were found to be between 95% and 115%, and the contamination was strongly reduced by using a number of precautions. Our results showed that the Rhone River was a source of PAE in estuary and in coastal area. Although DEHP was predominant in the river, its relative abundance decreased with increasing salinities, DiBP being predominant in the plume for a salinity of 28. We reported, for the first time, the PAE vertical profile (0-1500 m) from the offshore water, which had lower concentrations compared to the Bay of Marseilles. We observed that DEHP was also predominant in surface marine waters including offshore (0-250 m) and below the surface in Marseille Bay (5 m). The opposite trend was found for DnBP for which concentrations significantly increased in mesopelagic waters. High concentrations of DnBP, DEHP and DiBP were detected close to the bottom (30 m) of the Bay of Marseilles. Such feature might be related to diffusion of PAEs released from the sediment surface very likely induced from plastic debris degradation and previously accumulated in the bottom. Precise origin of such material still needs to be determined. Other processes such as photochemical oxidation reactions in surface waters, bacterial degradation, are certainly important processes that regulate as well the distribution of PAEs in the water column. High PAE concentrations found in coastal area near Marseilles city strongly suggest that these organic compounds are very likely abundant in the whole Mediterranean that is surrounded by urbanized areas and that is submitted to relatively high inputs of rivers and atmospheric deposition compared to the global Ocean.

Fig. 1 .

 1 Fig. 1. Area of study. The map shows the sampling stations: SUNMEX in the Bay of Marseilles (October 2014), offshore 'Antares station' (March 2015), Arles in Rhone River and the river plume (March 2014).

Fig. 2 .

 2 Fig. 2. Recovery of PAEs obtained in two elution fractions carried out with five different organic solvents (F1) and dichloromethane (F2).

Fig. 3 .

 3 Fig. 3. Typical capillary gas chromatogram of PAEs (DMP, DEP, DiBP, DnBP, BzBP and DEHP) isolated from the seawater sample (grey line) collected from the Marseilles Bay (northwestern Mediterranean Sea) in October 2014 and typical PAEs standard solution chromatogram (black line, 0.1 mg/L). The identification of PAEs was confirmed by GC/ MS. The mass spectra of the PAEs (not presented here) were in accordance with those published in the literature. DPP and DnOP were not detected in this sample.

Fig. 4 .

 4 Fig. 4. PAE relative abundances (%) detected in different Mediterranean Sea samples: Surface Rhone River and associated Plume (March 2014), Bay of Marseilles (October 2014) and surface 'Cortiou' station (November 2014).

Fig. 5 .Fig. 6 .

 56 Fig. 5. PAE relative abundances (%) detected in offshore Antares station-samples collected at the surface and at depths of 5, 50, 100, 250, 500, 1000 and 1500 m in March 2015.

Table 1

 1 Retention time and ions for selected ion monitoring (SIM) of PAEs by GC/MS.

	Compound	Molar mass (g mol À1 )	Retention time (min)	Quantifier ion (abundance)	Qualifier ion (abundance)
	DMP	194.2	5.55	163 (100)	194, 135, 77
	DEP	222.2	6.12	149 (100)	222, 121
	DPP	250.0	6.84	149 (100)	209, 191
	DiBP	278.4	7.22	149 (100)	223, 205
	DnBP	278.4	7.65	149 (100)	223, 205
	BzBP	312.4	9.72	149 (100)	206, 238
	DEHP	390.6	10.62	149 (100)	279, 167
	DnOP	390.6	11.76	149 (100)	279
	DEP-d4	226.2	6.11	153 (100)	181
	DnBP-d4	282.4	7.64	153 (100)	209
	DnBP-d4	282.4	7.64	153 (100)	209
	DEHP-d4	394.6	10.61	153 (100)	171, 283

Table 2

 2 Recoveries of PAE extractions obtained with 4 different volumes of spiked seawater and 200 and 400 mg of HLB sorbent.

	Compounds	Recoveries (%)
		200 mg				400 mg
		100 mL	200 mL	500 mL	1000 mL	100 mL	200 mL	500 mL	1000 mL
	DMP	99	108	100	105	114	107	109	110
	DEP	61	93	88	89	95	93	95	100
	DPP	105	117	113	120	122	121	119	114
	DiBP	96	122	111	123	121	100	107	115
	DnBP	79	75	63	66	136	77	67	82
	BzBP	102	125	94	106	126	104	87	97
	DEHP	14	13	8	10	11	7	21	6
	DnOP	8	10	15	10	10	6	24	6

Table 3

 3 Statistical data of the linear regression between the spiked concentration of PAEs and the measured peak areas. PAE recovery was reported.

	Compound	Spiked concentration (ng/L)	n	R 2	Slope (mVs/ng/L)	y-intercept (mVs)	Recovery (%)	MDL (ng/L)
	DMP	0, 5, 10, 20, 40	5	0.98	11,022	20,671	101.2	0.11
	DEP	0, 5, 10, 20, 40, 100	6	0.99	13,978	40,049	98.1	0.21
	DPP	0, 5, 10, 20, 40	5	0.98	12,072	2893	115.5	0.31
	DiBP	0, 40, 100, 200, 400, 600, 800	7	0.99	11,006	5,477,511	110.4	0.23
	DnBP	0, 40, 100, 200, 400, 600	6	0.99	19,424	1,965,660	99.7	0.28
	BzBP	0, 5, 10, 20, 40	5	0.99	2718	1845	98.7	1.67
	DEHP	0, 40, 100, 200, 400, 600, 800	7	0.99	1638	23,076	103.8	0.93
	DnOP	0, 5, 10, 20, 40	5	0.96	268	323	95.0	2.73

Table 4

 4 PAEs concentrations detected in different Mediterranean Sea samples: Surface Rhone River and associated Plume (March 2014), Bay of Marseilles (October 2014) and surface 'Cortiou' station (November 2014).

	Compounds	Rhone River	Rhone River plume		Bay of Marseilles			Cortiou
		S = 0	S = 10	S = 18	S = 28	1.5 m	5 m	15 m	30 m
	DMP	5.7	5.7	4.8	2.7	1.8	1.7	1.4	6.4	7.3
	DEP	30.5	29.0	30.2	16.4	12.7	6.9	7.1	50.0	33.5
	DPP	n.d.	n.d.	n.d.	n.d.	n.d.	n.d.	n.d.	n.d.	n.d.
	DiBP	125.9	48.6	67.0	136.4	215.7	56.5	77.0	383.4	81.7
	DnBP	40.5	21.8	21.7	22.8	87.0	63.4	138.9	466.0	28.8
	BzBP	5.4	2.7	2.6	2.5	3.5	3.2	4.0	4.8	5.1
	DEHP	406.8	101.2	55.9	39.2	103.5	112.8	102.9	296.5	56.2
	DnOP	n.d.	n.d.	n.d.	n.d.	n.d.	n.d.	n.d.	n.d.	n.d.
	Total	615.1	209.0	182.2	220.0	424.2	244.5	331.3	1207.1	212.6
	Table 5
	PAE relative abundances (%) detected in different Mediterranean Sea samples: Surface Rhone River and associated Plume (March 2014), Bay of Marseilles (October 2014) and
	surface 'Cortiou' station (November 2014).
	Compound	Rhone River	Rhone River plume		Bay of Marseilles			Cortiou
		S = 0	S = 10	S = 18	S = 28	1.5 m	5 m	15 m	30 m
	DMP	0.9	2.7	2.6	1.2	0.4	0.7	0.4	0.5	3.4
	DEP	5.0	13.9	16.6	7.5	2.8	2.8	2.1	4.1	15.7
	DPP	n.d.	n.d.	n.d.	n.d.	n.d.	n.d.	n.d.	n.d.	n.d.
	DiBP	20.5	23.3	36.8	62.0	47.3	23.1	23.3	31.8	38.4
	DnBP	6.6	10.4	11.9	10.4	19.1	25.9	42.0	38.6	13.5
	BzBP	0.9	1.3	1.4	1.1	0.8	1.3	1.2	0.4	2.4
	DEHP	66.1	48.4	30.7	17.8	22.7	46.0	31.1	24.6	26.4
	DnOP	n.d.	n.d.	n.d.	n.d.	n.d.	n.d.	n.d.	n.d.	n.d.

Table 6

 6 PAE relative abundances (%) detected in offshore Antares station-samples collected at the surface and at depths of 5, 50, 100, 250, 500, 1000 and 1500 m in March 2015.

	Compounds	Depth (m)
		0	5	50	100	250	500	1000	1500
	DMP	5.7	5.1	5.0	3.5	7.9	4.6	3.5	2.3
	DEP	4.4	2.3	2.0	1.9	2.3	1.6	1.7	1.5
	DiBP	7.9	5.1	4.5	3.5	4.4	3.6	4.5	3.5
	DnBP	15.2	10.9	15.2	25.8	28.5	57.0	61.0	72.6
	DEHP	66.9	76.7	73.3	65.4	56.9	33.1	29.2	20.1

Acknowledgments

This study was conducted as part of the PARTICULE-Région PACA, MERMEX-MISTRALS, EC2CO-PLASTIC-MICRO of M I O, LABEX OT-Med projects supported by Pole Mer Méditerranée. This study is a contribution to the international SOLAS and LOICZ projects. We acknowledge the financial support from the PACA region, which provided a PhD scholarship for A. Paluselli. The authors are grateful to the crews of the R/V Antedon II and SAM-M I O platform for their operation at sea. We also acknowledge Anne Delmont for help in seawater sampling and part of DOC analyses.


Jianglin Ye

Dr Huabing Tan

Shuilin Wu

Kun Ni

Fei Pan

Jie Liu

Prof Zhuchen H Tao

Yan Qu

Hengxing Ji

Patrice Simon

Prof. Y H Ji

Zhu

Prof P Simon

Direct Laser Writing of Graphene Made from Chemical Vapor Deposition for Flexible, Integratable Micro-Supercapacitors with Ultrahigh Power Output

Keywords: direct laser writing, graphene, line-filtering, micro-supercapacitors, power density

ou non, émanant des établissements d'enseignement et de recherche français ou étrangers, des laboratoires publics ou privés.

Direct Laser Writing of Graphene Made from Chemical Vapor Deposition for Flexible, Integratable Micro-Supercapacitors with Ultrahigh Power Output

Jianglin Ye, Huabing Tan, Shuilin Wu, Kun Ni, Fei Pan, Jie Liu, Zhuchen Tao, Yan Qu, Hengxing Ji, Patrice Simon, and Yanwu Zhu* DOI: 10.1002/adma.201801384 lithium-related safety issues. [2] As an alternative to microbatteries, the micro-supercapacitor (MSC) represents an emerging family of miniaturized electrochemical energy storage devices with intriguing performances of high power per volume, outstanding rate capability, extremely long lifetime, and mechanical stability. [3,4] Solid-state MSCs can be directly integrated with other miniaturized electronic devices, such as microbatteries and energy-harvesting microsystems. [5] Notable efforts have been devoted to improve the energy and power performances of MSCs via fabrication of newly developed electrode materials (e.g., onion-like carbons, carbide-derived carbons, doped-carbons, conducting polymers-coated carbons), [6,7] or by utilizing advanced thin-film preparation/processing techniques (e.g., layerby-layer assembly, roll-to-roll transfer/ printing, spray-coating, lithography, laserassisted method). [8][9][10][11] Nevertheless, the overall merits such as scalability, transferability, integrability, flexibility, electrochemical performance, and fabrication efficiency of MSCs are still far from satisfactory. [3] With high theoretical areal capacitance and other merits, graphene offers great potential for electrochemical energy storage. [12] Conventional supercapacitors have been fabricated with graphene-based or graphene-derived materials for improved performances. [13] Graphene-based film electrodes Development of high-performance yet flexible energy storage devices is essential to meet the growing needs for portable, wearable, and implantable electronic devices. [1] Currently portable devices mainly rely on batteries as micro power sources; lithium thin-film batteries (LTF) are commercially available. Such devices, however, suffer from relatively low power densities (typically <10 mW cm -3 ), limited cycle times, and have also been fabricated with high power density due to the rapid transport of electrolyte ions between the 2D platelets. [10,14,15] To fabricate graphene-based MSCs, spin-coating, vacuum filtration, and laser-assisted methods have been utilized. [16,17] Among others, inkjet printing and screen printing show simple and cost-effectiveness for efficient p atterning microelectrodes on various substrates, for which a stable ink with good compatibility to the target substrate is a stringent requirement. [9,14] The lithographic fabrication is an effective way to fabricate MSCs with high-resolution patterns, while involving many separated steps and high cost, to some extent, lacking of versatility. [18] In addition, most graphene-based film electrodes reported so far are based on reduced graphene oxide (rGO); the reduction processes, e.g., by high temperature, plasma or reductant treatment, however, are not suitable for scalable and integrated applications, and suffer from the restacking of graphene platelets as a result of the van der Waals interaction between platelets. [18,19] Wet processing may also bring problems in the fabrication of semiconductor devices, hampering the manufacturing of MSCs on silicon chips. [20] Recently, a photochemical reduction of GO with TiO 2 was developed for the scalable fabrication of graphene-based MSCs with excellent performances, in which masking is relied to define microelectrodes on substrates and the versatility of making different patterns on one chip may be sacrificed. [16] In terms of efficiency and substrate compatibility, direct laser writing (DLW) technique offers a promising protocol for fabrication of MSC arrays, both on flexible and rigid substrates, at the room temperature and over large areas. [21] For example, flexible solid-state MSCs have been demonstrated by DLW of polyimide films and MXenes, with a typical power density of <15 W cm -3 . [22,23] DLW has also been utilized to fabricate flexible MSCs by laser-assisted reduction of GO fi lms with a standard LightScribe digital video disc (DVD) drive, showing a power density of 200 W cm -3 . [4,24] But the low packing density of rGO films made by laser scribing has resulted in an uncompetitive capacitance, energy, or power density (typically, ≈8.7 F cm -3 , 0.98 mWh cm -3 in poly(vinyl alcohol) (PVA)/LiCl). High-performance MSCs, ideally in compact volume and with high power, are capable of alternating current (AC) line-filtering, while with designable energy density MSCs are desired for wearable, flexible electronic devices on versatile substrates.

In this work, we demonstrate that DLW is a highly efficient approach to fabricate solid-state MSCs based on graphene films made by mass chemical vapor deposition (CVD). Flexible MSCs prepared on various substrates with customer-designed patterns or integrations have shown superior and devisable performances, in both hydrogel and ionogel electrolytes. The solidstate MSCs based on 8 layers of CVD graphene deliver high specific capacitances of 62.7 µF cm -2 and 36.8 F cm -3 in PVA/ H 2 SO 4 hydrogel electrolyte. The short resistance-capacitance (RC) time constant of 0.54 ms of such multilayered graphene-MSCs (MG-MSCs) allows an excellent AC line-filtering performance which can be readily utilized in the conversion of AC signal to DC output. In an ionogel electrolyte, more remarkably, MG demonstrates high capacitance of 45.2 µF cm -2 (147 F cm -3 in electrode) at 10 mV s -1 , excellent rate capability indicated by a capacitance of 5.1 µF cm -2 at 500 V s -1 , ultrahigh volumetric power density of 1860 W cm -3 , and energy density of 23 mWh cm -3 ; both the power density and energy density are among the highest values achieved for carbon-based MSCs. Furthermore, by depositing ≈245 nm thick polyaniline (PANI) on graphene films followed by DLW, the specific capacitance of hybrid MSCs has been optimized to 3.8 mF cm -2 in PVA/ H 2 SO 4 hydrogel electrolyte, based on which an integrated line filtering/energy storage/sensing device is fabricated, showing the excellent integrability and scalability of DLW for the fabrication of MSC-based devices.

Figure 1A illustrates the fabrication procedure of MG-MSCs. Monolayer graphene (G) films grown on copper (Cu) foils with a typical size of A4 size (21 × 29.7 cm 2 ) are obtained from mass CVD production. Poly(ethylene terephthalate) (PET) pressuresensitive adhesive (PSA) films are used to assist the transfer of graphene on target substrates. [25] By repeating the dry transfer sequentially, it is ready to obtain n-layered CVD graphene films in A4 size on target substrates such as PET (Figure S1, Supporting Information). [8,26] Raman spectroscopy reveals that the intensities of G-band and 2D-band increase with the layer number of graphene transferred, while the G/2D ratio does not significantly change (Figure S2, Supporting Information), suggesting that each layer in the stacking remains the feature of monolayer graphene. [27] However, non-negligible D peaks obtained from the MG films with more layers indicate the higher content of defects when compared to fewer layer MG films. Compared to the monolayer graphene, the blue-shifted 2D peak in MG films may imply a slight p-doping which may be introduced during the sequential transfer. X-ray photoelectron spectroscopy (XPS) (Figure S3, Supporting Information) and sheet resistance measurements (Figure S4, Supporting Information) of MG films confirm the high carbon purity and the sequential reduction of sheet resistance with the graphene stacking layers. As the Au film is considered as an effective adhesive-strained layer to graphene, [28] MG films have been transferred to the target substrates deposited with Au films, followed by a DLW equipped with a computer-assisted design (CAD) for manufacturing microelectrode patterns. Nanometerhigh ripples are observed from the atomic force microscopy (AFM) images (shown in Figure S5, Supporting Information) of the MG films transferred on a PET/Au substrate, due to the wrinkles of CVD graphene. [29] With DLW, micropatterns with high resolution and high density are fabricated in a highly efficient manner. For example, more than 100 microdevices (in an area of about 1 × 1 cm 2 ) are produced on a single run in 5 min; the micropatterns on PET substrate with a size of ≈500 µm are completely flexible and can be attached to a human finger (Figure S6, Supporting Information). Figure 1B shows the micropattern arrays consisting of various designed patterns (further magnified and shown in Figure 1C) on one PET/Au substrate with a size of 10 × 10 cm 2 . In this way, devices with desirable patterns can be made on substrates such as polyimide (PI), Si/SiO 2 , and poly(tetrafluoroethylene) (PTFE) coated with Au films in addition to PET. Figure 1D shows a pattern constructed with 50 interdigitated microelectrodes on PET/ Au substrate. The ultrathin MG-MSCs on Teflon film can be twisted, further indicating the high flexibility (Figure 1E). The optical image (Figure 1F) shows the typical microelectrode width of ≈180 µm; a channel width of ≈70 µm is observed from scanning electron microscopy (SEM) (Figure 1G). Elemental mapping analysis confirms the uniform distribution of C element on the interdigital fingers and the clear space between fingers (Figure 1H). One shall notice that the width of microelectrodes and channels can be readily tuned in DLW by CAD.

The MG microelectrodes made by DLW are directly used as MSCs (MG-MSCs) after an electrolyte, e.g., 1-ethyl-3-methylimidazolium-bis (trifluoromethylsulfonyl) imide (EMI TFSI) ionic liquid, PVA/H 2 SO 4 hydrogel or fumed silica nanopowder/ ionic-liquid ionogel (FS/IL, e.g., fumed silica nanopowder/ EMI TFSI), is cast on the arrays. The electrochemical measurements of MG-MSCs in EMI TFSI electrolyte (Figure S7a, Supporting Information) show that the capacitance increases with the number of graphene layers in the stacking in a nearly linear manner, till ≈0.1 mF cm -2 for a device with 25-layered graphene films. To balance the performance and transfer cycles, MSCs with 8-layered graphene films were prepared on PET/Au and the performance was investigated systematically. From the cyclic voltammetry (CV) curves shown in Figure 2A,B, we can see that the MG-MSCs exhibit good electrochemical behavior with a nearly rectangular CV shape, demonstrating a typical behavior of double-layer capacitors at conventional (10 V s -1 ) and high scan rates (500 V s -1 ) in ionic gel electrolytes. Specifically, MG-MSCs are comparable to the ionogel electrolyte till an electrochemical window of 3.0 V (Figure S7b, Supporting Information). With the increase in the scan rate, the discharge current density (at 0 V) responds linearly to the scan rate till 100 V s -1 for both electrolytes (Figure 2C), reflecting the fast ion diffusion. The areal and volumetric capacitances of the MG-MSCs (Figure 2D) in PVA/H 2 SO 4 hydrogel electrolyte are 62.7 µF cm -2 (338 µF cm -2 in MG electrode) and 36.8 F cm -3 (197 F cm -3 in MG electrode) at the low scan rate, e.g., 10 mV s -1 . CVs recorded at lower scan rates show the contribution of redox process [32] with the faradic efficiency closing to 1 (Figure S8, Supporting Information). The comparison in Table S1 (Supporting Information) shows that the volumetric capacitance of MG-MSCs in PVA/H 2 SO 4 is higher than those reported from MSCs based on methane plasma rGO (MPG-MSCs, 17.5 F cm -3 ), [19] graphene/carbon nanotube carpets (G/CNTCs-MCs, 1.08 F cm -3 ), [33] laser-scribed graphene (LSG, 19 F cm -3 ), [4] and laser-processed graphene (LPG-MPS, 8.7 F cm -3 ). [34] Notably, MG-MSCs still deliver an areal capacitance of 5.9 µF cm -2 and a volumetric capacitance of 3.4 F cm -3 , at an ultrafast scan rate of 1000 V s -1 in PVA/H 2 SO 4 hydrogel electrolyte. In FS/IL ionogel electrolyte, MG-MSCs show specific capacitances of 45.2 µF cm -2 (251 µF cm -2 in MG electrode) and 26.5 F cm -3 (147 F cm -3 in MG electrode) at 10 mV s -1 , and highrate performance with 5.1 µF cm -2 and 3.0 F cm -3 , respectively, at 500 V s -1 , which are higher than the values reported in the literatures while obtained from much lower scan rates (5-200 V s -1 ), e.g., for onion-like carbon (0.4 F cm -3 ), LSG (1.4 F cm -3 ), and laser-induced porous graphene films (≈0.66 F cm -3 ) (also seen in Table S1, Supporting Information). [6,23] The cycling measurement (Figure S9a, Supporting Information) shows that the retention is ≈96% after 20 000 charge/discharge cycles at 100 V s -1 and the CV curves remain unchanged after cycling in hydrogel electrolyte. Meanwhile, MG-MSCs retain ≈82% of initial capacitance after 20 000 cycles at a scan rate of 10 V s -1 in FS/IL electrolyte. To examine the mechanical robustness, the MG-MSCs were experienced by repeated bending and the electrochemical results obtained after every 100 bending cycles (shown in Figure S9b, Supporting Information) reveal that ≈85% of the initial performance is remained after 4750 cycles (in PVA/H 2 SO 4 ) and 1900 cycles (in FS/IL), respectively. The flexibility of MG-MSCs was also evaluated by performing the electrochemical measurement and bending simultaneously; the CV curves in Figure S9c,d (Supporting Information) show that nearly 100% of capacitance is maintained for the bending of 180°, in both hydrogel and ionogel electrolytes. The comparison of the MG-MSC design on different substrates (Figure S10, Supporting Information) indicates that the difference in capacitance is less than 10% among the substrates. The device performances are also stable over several months in ambient conditions (Figure S11, Supporting Information). The Ragone plot shown in Figure 2E compares the performance of MG-MSCs with various micro energy storage devices. As can be seen, MG-MSCs exhibit exceptional energy density of 23 mWh cm -3 (at 0.32 W cm -3 ) in ionogel electrolyte (normalized to the MG electrodes), which is higher than those of other carbon-based thin-film MSCs (normalized to the carbon electrodes as well) such as doped graphene (sulfur-doped graphene, SG-MSCs, 12.4 mWh cm -3 ; nitrogen and boron codoped graphene, BNG-MSCs, 16.9 mWh cm -3 ), [30] MPG-MSCs (7.5 mWh cm -3 ), [19] photochemical reduction of GO/TiO 2 composite (PRG-MSCs, ≈13.2 mWh cm -3 ), [16] and ultrathin printable MSCs (G/PH1000, UPSCs-25, 12 mWh cm -3 ) [14] (see Table S1, Supporting Information), while lower than the ultrathin rGO film (32 mWh cm -3 ) obtained by hydrazine monohydrate reduction. [18] Simultaneously, MG-MSCs exhibit ultrahigh power densities of 1714 W cm -3 (29 mW cm -2 in PVA/H 2 SO 4 ) and 1860 W cm -3 (32 mW cm -2 in ionogel electrolyte). The ultrahigh power performance in ionogel electrolyte is superior to the state-of-the-art of high-power MSCs based on graphene, doped-graphene (typically lower than 1000 W cm -3 ), CNT/G composites (300 W cm -3 ), and nanostructured-MSCs (typically, 200-1400 W cm -3 ). [18,20,31] In terms of energy storage performance, MG-MSCs actually are superior to ultrahighpower electrolytic capacitors (>1000 W cm -3 ).

To obtain the hint about the mechanism for the ultrahigh power of MG-MSCs, the interlayer spacing of MG was estimated from the height measurement with AFM. From Figure S12 (Supporting Information), we can see that the interlayer distance remains about 1.2 nm during the sequential transfer of graphene films, which is much larger than the CV curves obtained at scan rates of A) 10 V s -1 and B) 500 V s -1 , respectively. C) Plots of discharge current as a function of scan rate. D) Evolution of areal and volumetric capacitances with scan rate. E) Ragone plots of MG-MSCs, compared with commercial LTF batteries, electrolytic capacitors, and carbon based MSCs such as SG-MSCs, [30] MPG-MSCs, [19] and PiCBA MSCs, [31] among others listed in Table S1 (Supporting Information). F) Schematic showing the layered geometry of MG microelectrodes suitable for the efficient ion storage and transport.

typical interlayer distance in rGO (e.g., ≈3.4 Å in high-temperature-reduced GO films, [19] or 5-7 Å obtained from hydrazine assisted reduction, electrophoretic deposition, or nanoparticles supporting [15,16,18] ). As schematically illustrated in Figure 2F, we propose that the 1.2 nm channels between the adjacent graphene layers offer feasibility for the highly efficient storage and superfast diffusion of electrolyte ions, which have the ionic sizes of, e.g., typically, ≈4 Å for SO 4 2-, PO 4 3-, HPO 4 2-, [START_REF] Speight | Lange's Handbook of Chemistry[END_REF] ≈7.6 Å for EMI + , and ≈7.9 Å for TFSI -. [START_REF] Tsai | [END_REF] To see this point more clearly, the average capacitance (C ave ) of each side of graphene layer in MG has been obtained by normalizing the areal capacitance (C electrode ) with 2n (n is the number of graphene layers) for MG-MSCs made from 2, 4, and 8 layers and is listed in Table S2 (Supporting Information). From Table S2 (Supporting Information), we can see that the specific capacitance normalized to one side of monolayer graphene gradually decreases with the layer number of MG electrodes, but higher than the double-layer capacitance of graphene (e.g., 21 µF cm -2 ) mostly reported. [37] This result could be explained by the extra capacitance contribution from the edges of MG, the doping and defects in MG, and the Au collector below MG, [38] in addition to that from the graphitic planes in MG (as a reference, the capacitance of the bare Au collector was measured as ≈10 µF cm -2 in PVA/H 2 SO 4 or ≈7 µF cm -2 in FS/IL at 10 mV s -1 , respectively). Furthermore, by assuming a voltage drop of 1.25 V on each electrode for a symmetric MSC in ionogel electrolyte, the charge density on each side of monolayer graphene in the 8-lay-ered MG is estimated as ≈16 µF cm -2 × 1.25 V = 0.012 e Å -2 , which is still lower than the maximum charge density of 0.031 or 0.044 e Å -2 , estimated from a close packing of anion (TFSI -) or cation (EMI + ), respectively. [39] The result suggests that the charge storage capacity in MG electrodes could be further optimized by further improving the accessible area, e.g., in EMI TFSI electrolyte, which shows a small contact angle of ≈36° on MG (Figure S13, Supporting Information).

Electrochemical impedance spectroscopy (EIS) has been used to further investigate the frequency response of MG-MSCs on PET/Au in PVA/H 2 SO 4 hydrogel and FS/IL ionogel electrolytes. The Nyquist plots shown in Figure 3A are typical from capacitive, blocking electrodes and an equivalent series resistance of 9.8 Ω in PVA/H 2 SO 4 electrolyte or 23.1 Ω in FS/IL electrolyte is obtained. From the curves of phase angle versus frequency (Figure 3B), the characteristic frequency f 0 at -45° is 2025 Hz (in PVA/H 2 SO 4 ) and 1150 Hz (in FS/IL), corresponding to a time constant (τ 0 ) of 0.49 and 0.87 ms, for PVA/H 2 SO 4 and FS/IL electrolyte, respectively. These values are lower than or comparable to those reported from ultrathin graphene-based MSCs in hydrogel electrolyte (Figure 3C), e.g., UPSCs-25 (G/PH000, 1 ms), [14] MPG-MSCs (0.28 ms), [19] SG-MSCs (0.26 ms), [30] azulene-bridged coordination polymer framework film (coordination of 2,2′-diisocyano-1,1′,3,3′-tetraethoxycarbonyl-6,6′-biazulene (donted as iCBA), PiCBA, 0.27 ms), [31] G/CNTCs-MCs (0.82 ms), [33] and PRG-MSCs (16.4 ms) (Table S1, Supporting Information), [16] indicating the efficient ion transport/electron conductance in MG microelectrodes. The phase angle at 120 Hz was measured to be -76.2° with a RC time constant (τ RC ) calculated from the imaginary capacitance of 0.54 ms in PVA/ H 2 SO 4 (Figure S14, Supporting Information), one order lower than Al electrolytic capacitor (AEC) (-83.4°, 8.3 ms), [40] demonstrating the potential of MG-MSCs for flexible AC line-filtering device. [41] Scheme S1 (Supporting Information) shows an AC line filtering electrical circuit composed of a full wave bridge rectifier, smoothing MG-MSCs and a load resistor. The input AC signal is generated by function generator (V peak = ±2.3 V, 50 Hz) and the full-bridge rectifier converts the AC signal to a pulsating DC signal (Figure 3D), which is converted to a smooth DC output by the MG-MSCs eventually. As can be seen in Figure 3E, single MG-MSC converts the rectified pulsating signal to a smooth DC output of ≈0.88 V and two MG-MSCs in series convert the signal to a stable output of ≈1.6 V. It is noted that there was no obvious change in DC output when the measurement was conducted under a bending of ≈60° and the output power can reach up to 11 mW (for two MG-MSCs in series with an area of ≈3 cm 2 ).

We demonstrate here that DLW is a universal protocol for the construction of an integrated energy harvesting device based on piezoelectric generator (PEG), MG-MSCs line-filtering, and graphene-polymer hybrid films/sensors (Figure 4A). We are also aware of the limitations of MG-MSCs in terms of areal capacitance or energy density (0.04 µWh cm -2 ) accompanying with the very small thickness, which make MG-MSCs more suitable to be used in high power AC line-filtering devices. On the other hand, the capacitance was also optimized by depositing pseudocapacitive active materials on MG films. As shown in Scheme S2 (Supporting Information), MG films (4-8 layers) were used as conductive substrates for electrochemical deposition of PANI films with a size of 12 × 12 cm 2 (Figure S15a, Supporting Information). The SEM image (Figure S15b, Supporting Information) shows that the 245 nm thick PANI (from AFM image in the Figure S15c, Supporting Information) exhibits a uniform surface with a rod-like morphology on the top of MG film. Based on the MG-PANI MSCs fabricated by DLW (Figure S15d, Supporting Information), the specific capacitance in PVA/H 2 SO 4 electrolyte is improved to 3.8 mF cm -2 and 155 F cm -3 in device at 1 mV s -1 (Figure S16, Supporting Information). The obtained MG-PANI MSCs demonstrate significantly improved energy densities of 21 mWh cm -3 (at 0.08 W cm -3 ) in PVA/H 2 SO 4 electrolyte and 34 mWh cm -3 (at 1.2 W cm -3 ) in poly(methyl methacrylate) (PMMA)/LiClO 4 electrolyte (Figure 4C; Figure S17 and Table S3, Supporting Information), comparable to or higher than those of state-of-the-art MSCs such as polypyrrole (PPy) nanowires, V 2 O 5 /exfoliated graphene MSCs using similar electrode thickness. [42] The rod-like, intertwined PANI morphology has also contributed to the ultraflexibility and exceptional mechanical stability of MG-PANI MSCs, as indicated by the high capacitance retention of 86% when the device is bent at an angle of 150° for 1000 cycles in PVA/H 2 SO 4 electrolyte (Figures S18 andS19, Supporting Information).

An integrated device, containing a diode bridge rectifier circuit, two MG-MSC line-filter devices in series, MG-PANI MSCs array (2 × 2) (both in PVA/H 2 SO 4 electrolyte), and pressure/gas sensors is fabricated in one step by DLW on PI/Au substrate (Figure 4B). The input AC signal is generated by pressing PEG at a frequency of 110 Hz, which is smoothed by the MG-MSC line-filter and eventually stored by MG-PANI MSCs. After charging the MG-MSCs, the DC voltage smoothly reaches ≈1.5 V (Figure 4D). The charged MG-MSCs and MG-PANI MSCs can be galvanostatically discharged at 4 µA cm -2 for ≈9 s and ≈500 s, respectively. From the current leakage evaluation of the devices (Figure 4E), we can see that the leakage current quickly drops to only 1.3 µA (0.4 µA cm -2 ) for MG-PANI MSCs and 120 nA for MG-MSCs, respectively, indicating the high stability of the MSCs. The pressure detector shows a steady response to pressure pulses (≈5 kPa), as shown in Figure 4F. The on/off ratio remains 10 000 after 10 000 s. The sensor with integrated MG-PANI MSCs successfully detected walking. As shown in Figure 4G, repeated measurements of each step give regular signals, signifying the reproducible sensing powered by the integrated MSC array and potential applications in wearable energy storage. The response current gradually decreases with time, which is attributed to the gradual decrease in the output voltage of the charged MG-PANI MSCs with time. In addition, PANI gas sensors respond quickly to toxic gas such as 200 ppm of NO 2 or NH 3 with high sensitivity and fast response at room temperature (22-25 °C) with the integrated device as power source; [43] the sensitivity tested also remains stable till 0 or 35 °C. As shown in Figure 4H, the different response to NO 2 or NH 3 is explained by the transfer of electrons or protons between PANI and gas molecules (inset of Figure 4H), [44] reflecting the feasibility of the wearable, self-powered integrated device.

In summary, we have demonstrated several prototypes of CVD-graphene-based in-plane MSCs with superior performances by DLW. The mass production and stable dry transfer of CVD graphene have guaranteed the highly stable interlayer distance of the multilayered graphene packing, benefiting to the efficient ion transport and simultaneously the excellent electric conductance of MG microelectrodes. The as-fabricated flexible MSCs delivered excellent specific volumetric power and energy densities and showed promising AC line-filtering performance with an extremely short RC time in PVA/H 2 SO 4 electrolyte.

It should be noted that the areal capacitances or energy densities of MG-MSCs were limited by the very small thickness of MG electrodes restricted by the tedious sequential transfer process, though which can be carried out in an automatic way. The hybrid films made by depositing electrochemically active materials, such as conductive polymers on MG demonstrated above, have provided further opportunities for optimizing the performance of the MSC device. Some other strategies such as chemical doping or fine-tuning of structure/morphology may further increase the energy/power density of MG-based MSCs. The DLW fabrication process has shown potential for scaling to large area and for integration with other devices. We believe this study opens up a plethora of versatile design and scalable fabrication of flexible, on-chip devices for practically useful energy storage devices based on CVD graphene.

Experimental Section

Fabrication of MG-MSCs: The monolayer graphene films were obtained by CVD on copper (Cu) substrates from Wuxi Graphene Film Co., Ltd. The layer-by-layer transfer of CVD graphene from Cu to target substrates was carried out with the assistance of PSA films, as previously reported with further modification. [25] Rollers (≈100 mm min -1 ) were used to assure the adhesion between the graphene film grown on Cu with the adhesive layer. Before transfer, a 25 nm gold (Au) film was first sputtered on the substrate with a rate of ≈1.0 Å s -1 in a chamber with a pressure of ≈2 × 10 -6 torr (Kert Lesker Lab18) with a size of up to A4. After sequential transfers, MG films were obtained on various target substrates (e.g., PI, PTFE, PET, Si/ SiO 2 ) coated with Au. For the fabrication of MG-MSCs, interdigital MG microelectrodes were created by utilizing direct ultraviolet laser machining under ambient conditions (Design Comes True, DCT). The laser power and scanning speed were set as ≈1 W and 200 mm s -1 , respectively. To fabricate MG-MSCs, PVA/H 2 SO 4 hydrogel electrolyte or FS/IL ionogel electrolyte was slowly casted onto the surface of MG microelectrodes and solidified for overnight. The hydrogel electrolyte was prepared by mixing 6 g of PVA and 6 g of sulfuric acid (H 2 SO 4 ) into 60 mL of deionized (DI) water, and heated at 80 °C for 3 h under stirring. The ionogel electrolyte was prepared by mixing fumed silica nanopowder (particle size of 7-40 nm) with ionic liquid (EMI TFSI) with weight ratio of 0.01 g/1 g, and then the mixture was stirred for 10 h under an Ar atmosphere to produce a clear viscous ionogel (named as FS/IL).

Fabrication of MG-PANI MSCs: PANI was electrochemically deposited on the MG films transferred on target substrates. For the deposition, the MG film was used as the working electrode in a three-electrode configuration that was constructed with a Pt plate as counter electrode and Ag/AgCl as reference electrode. The electrolyte was mixture of H 2 SO 4 (0.5 m) and aniline (0.05 m) in H 2 O. The electropolymerization of aniline on the MG film was performed at a constant potential of 0.75 V versus reference electrode for various durations. As-fabricated MG-PANI films were rinsed in DI water for several times and dried in a vacuum oven at room temperature. MG-PANI MSCs were fabricated by sequentially performing DLW of MG-PANI films and electrolyte casting following the same procedures as described above.

Characterizations and Electrochemical Evaluations: Material characterizations were performed by SEM (JEOL-6700F, Japan), AFM (Veeco Nanoscope Icon, US), Raman (Renishaw, 532 nm laser, UK), XPS (Thermo ESCALAB 250, US) with Al Kα radiation (hv = 1486.6 eV), and contact angle measurement (Dataphysics, OCA20, Germany). The electrochemical properties were examined by CV, galvanostatic charge and discharge, and EIS recorded in the frequency range of 0.1 Hz-100 kHz with a 5 mV AC amplitude, on a PARSTAT MC electrochemical workstation (AMETEK, US). An AC line filtering circuit was composed of four diodes (1N4007), a resistor, and MG-MSCs made on PI substrate. The output signal was measured using an oscilloscope (Tektronix, TDS2012C).

High-performance yet

  flexible micro-supercapacitors (MSCs) hold great promise as miniaturized power sources for increasing demand of integrated electronic devices. Herein, this study demonstrates a scalable fabrication of multilayered graphene-based MSCs (MG-MSCs), by direct laser writing (DLW) of stacked graphene films made from industry-scale chemical vapor deposition (CVD). Combining the dry transfer of multilayered CVD graphene films, DLW allows a highly efficient fabrication of large-areal MSCs with exceptional flexibility, diverse planar geometry, and capability of customerdesigned integration. The MG-MSCs exhibit simultaneously ultrahigh energy density of 23 mWh cm -3 and power density of 1860 W cm -3 in an ionogel electrolyte. Notably, such MG-MSCs demonstrate an outstanding flexible alternating current line-filtering performance in poly(vinyl alcohol) (PVA)/ H 2 SO 4 hydrogel electrolyte, indicated by a phase angle of -76.2° at 120 Hz and a resistance-capacitance constant of 0.54 ms, due to the efficient ion transport coupled with the excellent electric conductance of the planar MG microelectrodes. MG-polyaniline (MG-PANI) hybrid MSCs fabricated by DLW of MG-PANI hybrid films show an optimized capacitance of 3.8 mF cm -2 in PVA/H 2 SO 4 hydrogel electrolyte; an integrated device comprising MG-MSCs line filtering, MG-PANI MSCs, and pressure/gas sensors is demonstrated.

Figure 1 .

 1 Figure 1. Facile electrode design of solid-state flexible MG-MSCs using transferred CVD graphene films. A) Schematic illustrating the fabrication process of solid-state flexible MG-MSCs. B,C) Optical images of micropattern arrays with various geometries on PET substrates. Typical photograph of micropatterns on D) PET and E) more flexible PTFE films. F) Magnified optical image. G) SEM image of a MG channel and H) EDS carbon elemental mapping on Si/SiO 2 /Au.

Figure 2 .

 2 Figure 2. Electrochemical characterizations of MG-MSCs on PET substrates in PVA/H 2 SO 4 hydrogel and FS/IL ionogel electrolytes.CV curves obtained at scan rates of A) 10 V s -1 and B) 500 V s -1 , respectively. C) Plots of discharge current as a function of scan rate. D) Evolution of areal and volumetric capacitances with scan rate. E) Ragone plots of MG-MSCs, compared with commercial LTF batteries, electrolytic capacitors, and carbon based MSCs such as SG-MSCs,[30] MPG-MSCs,[19] and PiCBA MSCs,[31] among others listed in TableS1(Supporting Information). F) Schematic showing the layered geometry of MG microelectrodes suitable for the efficient ion storage and transport.

Figure 3 .

 3 Figure 3. High-frequency response of flexible AC-line filtering devices. A) Nyquist plots of MG-MSCs in PVA/H 2 SO 4 gel electrolyte (black) and FS/IL ionogel electrolyte (red). Inset shows the magnified plots in the high-frequency region. B) Impedance phase angle versus frequency. C) Radar plot of MG-MSCs, compared with Al electrolytic capacitor (AEC) and carbon-based MSCs. D) AC input signal and rectified pulsating DC signal. E) DC output of MG-MSCs. The inset shows that bending the device does not affect the charge storage.

Figure 4 .

 4 Figure 4. Design of flexible, integrated devices. A) Schematic and electrical circuit diagram of an integrated energy harvesting device made on PI. B) Photograph of the integrated energy harvesting/pressure/gas sensors device. C) Ragone plots for MG-PANI MSCs with comparison to reported energy storage devices in PVA/H 2 SO 4 electrolyte. D) Charging curve of the MG-MSCs by tapping the PEG and discharging curves of MG-MSCs and MG-PANI MSCs at the current of 4 µA cm -2 . E) Leakage currents of two MSCs over time. F-H) Response of current upon the detection of pressure (F), walking (G), and exposure to NO 2 and NH 3 (H).

Acknowledgements

This work was supported by the China Government 1000 Plan Talent Program, Natural Science Foundation of China (grant nos. 51322204 and 51772282). The authors thank the support from USTC Center for Micro-and Nanoscale Research and Fabrication.

Supporting Information

Supporting Information is available from the Wiley Online Library or from the author.

Conflict of Interest

The authors declare no conflict of interest.