C Caplat

O Basuyaux

S Pineau

J Deborde

A M Grolleau

Stéphane Le Glatin

Marie Laure Mahaut

Transfer of elements released by aluminum galvanic anodes in a marine sedimentary compartment after long-term monitoring in harbor and laboratory environments

Keywords: galvanic anode, aluminum, zinc, marine sediments, harbor, quality index 1. Introduction

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

2

Numerous maritime structures are constituted by metallic components (sheet piles in harbor environment, monopiles foundation of offshore wind turbines and platforms or pipelines). This type of structure undergoes a normal process of corrosion when exposed in the marine environment (Kaesche, 2003). A study by NACE International (2002) estimates that the total cost of marine corrosion in the world ranges from $ 50 billion to $ 80 billion a year. The corrosion process of a metallic material can be limited by different methods of protection such as coatings, corrosion inhibitors, corrosion allowance (increasing the thickness of material) or anodic and cathodic protections [START_REF] Kruger | Corrosion of Metals: Overview[END_REF][START_REF] Kirchgeorg | Emissions from corrosion protection systems of offshore wind farms: Evaluation of the potential impact on the marine environment[END_REF].

The methods of "cathodic protection" (CP) are provided by two mainly used electrochemical devices: cathodic protection by galvanic anodes (GACP) and cathodic protection by impressed current (ICCP) (DNV-RP-B401, 2017). In marine environments, the ICCP is rarely used, while CP by galvanic anodes is preferred for several reasons such as easy installation and less maintenance [START_REF] Roy | Aluminium alloy anode and method of using same in cathodic protection[END_REF]. The GACP consists of coupling two metallic materials that have a free and distinct potential.

Then, the coupling of both materials induces an electrochemical cell and decreases the potential of the noblest material. Reactions of oxidation occur with the consumption of these anodes and induce the liberation in waters of the elements that compose the anodes.

Different metals, such as zinc (Zn), magnesium (Mg) or aluminum (Al), are principal compounds of galvanic anodes. In a marine harbor environment, the use of Al anodes (at least 93.2% given by supplier) has been well known since the 1960s as a protectant for piles and sheet piles [START_REF] Roy | Aluminium alloy anode and method of using same in cathodic protection[END_REF][START_REF] Barbucci | Activation of aluminium anodes by the presence of intermetallic compounds[END_REF][START_REF] Shibli | Electrochemical impedance spectroscopic analysis of activation of Al-Zn alloy sacrificial anode by RuO2 catalytic coating[END_REF]. One of the principal reasons is the theoretical efficiency of this element in relation to its high electrochemical capacity (2000 Ah kg -1 ) in seawater compared to Zn which usually composes the galvanic anode used in marine environment (DNV-RP-B401, 2017; [START_REF] Kirchgeorg | Emissions from corrosion protection systems of offshore wind farms: Evaluation of the potential impact on the marine environment[END_REF]. Al is one of the most abundant elements on Earth after oxygen and silicon and the first metal composing the Earth's crust (7.5% in mass). This metal is widely distributed in soils and rocks where it naturally exists under different chemical forms, such as oxides or aluminosilicates in feldspars (major constituents of granite and silicon-rich rocks), in mica or in clays (ATSDR, 2008). Al inputs to the surface waters are largely associated with mineral particulate matter [START_REF] Driscoll | The chemistry of aluminum in the environment[END_REF], which quickly accumulate in sediments in which the anthropic part appears to be difficult to assess. In seawater (pH 8.2), dissolved inorganic aluminum is principally met in aluminate (Al(OH)4 -, 68%) and non-colloidal neutral hydroxide species (Al(OH)3 0 , 32%) [START_REF] Millero | Effect of Ocean Acidification on the Speciation of Metals in Seawater[END_REF]. Another important factor that increases the solubility of Al is the concentration of organic matter. In the presence of highly dissolved organic matter, Al solubility grows in relation to complexation with organic ligands [START_REF] Driscoll | The chemistry of aluminum in the environment[END_REF][START_REF] Ščančar | Aluminium speciation in environmental samples: a review[END_REF]ATSDR, 2008;[START_REF] Hagvall | Speciation of aluminum in soils and stream waters: The importance of organic matter[END_REF].

The speciation of Al is complex and determines its toxicity to organisms. A number of works have been carried out on Al toxicity in marine environment but were mostly associated with the inputs of acidic rivers into the coastal waters [START_REF] Wilson | Toxicity of Acid-Sulfate Soil Leachate and Aluminum to Embryos of the Sydney Rock Oyster[END_REF][START_REF] Kroglund | Low concentrations of inorganic monomeric aluminum impair physiological status and marine survival of Atlantic salmon[END_REF][START_REF] Golding | Derivation of a water quality guideline for aluminium in marine waters: Derivation of a marine water quality guideline for aluminium[END_REF][START_REF] Angel | Geochemical controls on aluminium concentrations in coastal waters[END_REF][START_REF] Gillmore | Toxicity of dissolved and precipitated aluminium to marine diatoms[END_REF]. The first studies about the impact of elements released from galvanic anodes were performed on anodes principally composed by Zn (at least 99% given by supplier) [START_REF] Bird | Zinc inputs to coastal waters from sacrificial anodes[END_REF][START_REF] Rousseau | Cathodic protection by zinc sacrificial anodes: Impact on marine sediment metallic contamination[END_REF][START_REF] Mottin | Effect of zinc sacrificial anode degradation on the defence system of the Pacific oyster, Crassostrea gigas: Chronic and acute exposures[END_REF]. Zinc is present in seawaters under ionic (including labile species), inorganic complexes and colloids as well as weak organic complexes [START_REF] Fukai | Chemical Forms of Zinc in Sea Water[END_REF]. This is an essential trace metal in living organisms [START_REF] Dušanka | Methodology of zinc determinations and the role of zinc in biochemical processes[END_REF] which tends to be bioaccumulated [START_REF] Pempkowiak | Speciation of heavy metals in marine sediments vs their bioaccumulation by mussels[END_REF][START_REF] Baudrimont | Geochemical survey and metal bioaccumulation of three bivalve species (Crassostrea gigas, Cerastoderma edule, and Ruditapes philippinarum) in the Nord Me´doc salt marshes (Gironde estuary, France)[END_REF]. Its chronic exposure effects in the coastal marine environment have been previously investigated by several authors [START_REF] Watling | Comparative Study of the Effects of Zinc, Cadmium and Copper on the Larval Growth of Three Oyster Species[END_REF]Sunda et al., 1987;[START_REF] Geffard | Temporal variations of metallothionein and metal concentrations in the digestive gland of oysters (Crassostrea gigas) from a clean and a metalrich site[END_REF][START_REF] Geffard | Bioaccumulation of Metals in Sediment Elutriates and Their Effects on Growth, Condition Index, and Metallothionein Contents in Oyster Larvae[END_REF]Devos et al., 2012). In the same period and more recently, some works focused on the Al released from anodes into the different compartments of marine environment [START_REF] Mao | Assessment of sacrificial anode impact by aluminum accumulation in mussel Mytilus edulis: A large-scale laboratory test[END_REF][START_REF] Gabelle | The impact of aluminium sacrificial anodes on the marine environment: A case study[END_REF][START_REF] Deborde | Impact of Galvanic Anode Dissolution on Metal Trace Element Concentrations in Marine Waters[END_REF]. But, no works were yet involved on the bioaccumulation potential of aluminum released from anodes.

This study followed up on the works of [START_REF] Deborde | Impact of Galvanic Anode Dissolution on Metal Trace Element Concentrations in Marine Waters[END_REF], which were conducted to optimize the simulation in the laboratory of the Al anodes degradation and to characterize the deposits of corrosion formed on anode's surface. Here, investigations were focused on in situ monitoring of sediments exposed for a long period to Al GACP anodes in natural and laboratory conditions. The sedimentary compartment was chosen because is considered as an indicator of contamination, since it is recognized as being able to store many contaminants such as metals [START_REF] Xia | Eighty-year sedimentary record of heavy metal inputs in the intertidal sediments from the Nanliu River estuary, Beibu Gulf of South China Sea[END_REF][START_REF] Li | Integrated Assessment of Heavy Metal Contamination in Sediments from a Coastal Industrial Basin, NE China[END_REF][START_REF] Pan | Trace metal contamination in estuarine and coastal environments in China[END_REF].

The metal concentrations recorded in sediments were first compared to guidelines [START_REF] Long | Incidence of adverse biological effects within ranges of chemical concentrations in marine and estuarine sediments[END_REF][START_REF] Macdonald | Development and evaluation of sediment quality guidelines for Florida coastal waters[END_REF][START_REF] Long | Predicting toxicity in marine sediments with numerical sediment quality guidelines[END_REF][START_REF] Mccauley | Sediment quality guidelines and assessment: overview and research needs[END_REF][START_REF] Zahra | Enrichment and geo-accumulation of heavy metals and risk assessment of sediments of the Kurang Nallah-Feeding tributary of the Rawal Lake Reservoir[END_REF]. They were used to determine sediment quality indices in the aim to evaluate the degree of contamination of the exposed sediments and their potential ecological risk [START_REF] Pourabadehei | Effect of the resuspension technique on distribution of the heavy metals in sediment and suspended particulate matter[END_REF][START_REF] Jafarabadi | Spatial distribution, ecological and health risk assessment of heavy metals in marine surface sediments and coastal seawaters of fringing coral reefs of the Persian Gulf, Iran[END_REF][START_REF] Jahan | Comparison of pollution indices for the assessment of heavy metals in the sediments of seaports of NSW, Australia[END_REF][START_REF] Gargouri | Use of sediment quality indicators for heavy metals contamination and ecological risk assessment in urbanized coastal zones[END_REF][START_REF] Kim | Critical evaluation of different methods to calculate the Geoaccumulation Index for environmental studies: A new approach for Baixada Santista -Southeastern Brazil[END_REF][START_REF] Tunca | An ecological risk investigation of marine sediment from the northern Mediterranean coasts (Aegean Sea) using multiple methods of pollution determination[END_REF]. And a chemical speciation was investigated to provide information on mobility and potential bioavailability of the considered elements [START_REF] Morillo | Heavy metal distribution in marine sediments from the southwest coast of Spain[END_REF][START_REF] Gabelle | The impact of aluminium sacrificial anodes on the marine environment: A case study[END_REF].

Materials and methods

Al galvanic anodes (Al GACP)

All the Al galvanic anodes (Al Bera15, Bac Corrosion ® Control) settled in situ in the harbor or that were used during the laboratory experiment were in the same Al alloy, which was mainly composed of seven elements (Al, Zn, Fe, Si, In, Mn and Cu). The mass percentages of each metal were given by the supplier as following: 93.2% to 97.1% of Al, 2.5% to 6.5% of Zn, 0.13% max of Fe, 0.1% max of Si, 0.01% à 0.03% of In, 0.01% max of Mn and 0.006% max of Cu. Depending of the supplier, the Al galvanic anodes can present impurities such as Cd, whose the mass percentage is less than 0.002% (DNV-RP-B401, 2017).

In situ sampling

The in situ study was conducted between September 2012 and June 2017 (42 months) in the Port of Calais located in northern France. As the great majority of harbor contracting authorities worldwide, the Port of Calais chose the method of CP by galvanic anodes to protect its immerged structures.

Approximately 615 tons of galvanic anodes were installed in 2014 in the harbor to protect the metallic structures for a duration of at least 15 years. Fig. 1 shows the location of the three sampling points monitored throughout the port: point 1 (at the entrance of the port) was considered to be the reference point; point 2 was in agitated area with important sediment resuspension in relation to the comings and goings of ferries (every 45 min, 350 to 400 boats per month); point 3 was located in the closed basin of a marina, which was under non-agitated area characterized by weak hydrodynamic condition and weak renewal of seawater. Point 2 is the only studied point that was subjected to soft dredging operations (water injections) performed two times per year.

In the port, all along the quays, one anode (approximately 700 kg) was fixed in the immersed zone at every one meter directly in contact with the steel sheet pile wall, in each hollow part. The pose of the anodes was distributed with time and the monitoring of harbor sediments was divided into two periods: first period corresponding to one year study carried out before the pose of the anodes, between September 2012 and September 2013; and the second period was developed over 30 months (January 2015 to end of June 2017) after the pose of the anodes, which ended in November 2014 for point 2 and in August 2014 for point 3.

Sediment samples were collected at proximity and below the anode (0 m) and 15 m away from the anode (15 m) for points 2 and 3. Two samples of undisturbed surface sediment samples were collected every 2 months by boat with a Van-Veen Surficial (top 2 cm) grab. Each sample was constituted by one kilogram of sediment obtained after mixing of three homogeneous grabs taken from an area of approximately 1 m 2 . Sediments were transferred in polyethylene bottles and placed in an ice-cooled box until their return in the laboratory. At same time, natural populations of mussels were collected, as well as caged mussels (Mytilus edulis) and marine sponges (Hymeniacidon perlevis) that were previously collected in an uncontaminated area of Normandy. However, the results of in situ organism monitoring will be published in a second part of this work.

Laboratory experimental design

The laboratory experiment was conducted to reproduce the in situ degradation of galvanic anode. This experiment was intended to monitor the elements released from the anode in the different compartments of marine environment, in agitated and non-agitated conditions, such as in an in situ environment (harbor). Sediments and marine organisms (the mussel M. edulis and the sponge H. perlevis) were exposed to the release of these elements in two different tanks (60 L) connected to one to another by the dumping of seawater (Fig. 2). The results of organism monitoring will be published in a second part of this work.

A laboratory experiment was performed between September 2014 and March 2016 (18 months) in NAVAL GROUP Research buildings located at Cherbourg. For this experiment, natural sediments were collected near Port of Calais (free of contamination) to respect similar size distribution and physical-chemistry proprieties in comparison to in situ harbor sediments.

Sediments were placed in the bottom of 3 tanks of 15-cm thickness covered with 30 cm of water that was renewed with a water flow of 80 L h -1 . The 3 conditions of the experiment were the following: a tank without anode, which was considered to be the reference tank (tank 1); 2 tanks with one anode positioned in suspension below the surface of the water, one in an agitated condition through a water pump brewing 3.33 L min -1 (tank 2) and one in non-agitated condition only with water renewal (tank 3). Three small cores (3 cm of thickness, approximately 60 mL) of undisturbed surface sediment samples were collected every 2 months in each tank. Sediments were transferred in polyethylene bottles and were placed in an ice-cooled box until their return to the laboratory.

Ten days after the supplying of sediments in the tanks, the anodes were connected to metallic plates (in carbon steel) to reproduce cathodic protection of the metal as in the in situ condition of the port.

For that, the steel sample was proportioned to allow the coupled galvanic anode to discharge an average current of approximately 100 mA, following the first laboratory tests that were performed in the previous study [START_REF] Deborde | Impact of Galvanic Anode Dissolution on Metal Trace Element Concentrations in Marine Waters[END_REF].

Electrochemical measurements (intensity and potential) of anodes were raised every two months. The results gave a mean current intensity of 93 ± 19 mA and a mean potential of -1063 ± 29 mV for the anode placed under the agitated condition (tank 2) and a mean current intensity of 82 ± 28 mA and a mean potential of -1058 ± 21 mV for the anode positioned under the non-agitated condition (tank 3). 8

Sediment characterizations

Once they reached the laboratory, sediment samples were immediately transferred to a cold chamber (4 °C) before the analyses of the particle size distribution and the total organic matter (TOC). These physical characteristics were performed for three periods of both in situ harbor monitoring and laboratory experiment, at the beginning, the middle and the end of each experiment. A laser method was applied to characterize the distribution of the particle size of sediments in accordance to the ISO standard 13320. The ISO standard 14235 was used to obtain the TOC.

The second kilogram of sediment samples was frozen at -20 °C immediately upon returning to the laboratory to be freeze-dried in a Labconco Freezone 2.5-L system with the aim to carry out the element analyses. Total concentrations of elements constituting galvanic anode (Al, Cu, Fe, In, Mn, Si and Zn) were analysed every two months for each experiment (harbor and laboratory monitoring).

To determine the total metal concentrations, sediment samples were digested by the alkaline fusion method following the ISO standard 14869-2.

Sequential extraction method

The chemical distribution in different fractions of sediment (acid-soluble, reducible, oxidizable and residual) of elements constituting the galvanic anode was determined by a sequential extraction method that was proposed by the European Union's Standards, Measurements and testing program (SM&T, formerly BCR). A detailed description of the method is found in [START_REF] Rauret | Improvement of the BCR three step sequential extraction procedure prior to the certification of new sediment and soil reference materials[END_REF] and consisted of three steps of successive extractions that allow for the part of the elements that are associated with one of the following fractions to be known:

Acid-soluble fraction (fraction F1):

The metals associated with this fraction are specifically nonadsorbed to solid particles (exchangeable metals) or are precipitated with natural carbonated forms, which are sensitive to pH variation. This fraction is the most labile linked to the sediment and is considered to be the most dangerous for the environment.

Reducible fraction (fraction F2):

This fraction contains metals bound to iron and manganese oxides that may be released if the sediment changes from the oxic to the anoxic state.

Oxidizable fraction (fraction F3): This fraction includes metals linked to organic matter and sulfides that can be released under oxidizing conditions.

A fourth fraction, called residual fraction, was determined for each metal by the difference between their total concentration and the sum of the concentrations that were obtained in the three previous fractions (F1+F2+F3). The residual fraction represents the parts of metal that are associated with minerals contained in the crystalline structure of solid particles (natural fraction) and are not being removed from the sediment.

Analytical method

The analyses of elements (Al, Cu, Fe, In, Mn, Si, Zn) were carried out by ICP-AES (Vista MPX, Agilent Technologies ® Santa Clara, USA).

The accuracy of the analytical procedure for total metal determinations was checked using CRM

PACS-3 (Marine Sediment Reference Material for Trace Metals) available from INMS (Institute for

National Measurement Standards) of the NRC (National Research Council Canada). The relative errors that were calculated between the experimental and certified concentrations were satisfactory for environmental analysis with values lower than 9% for all concentrations. Another certified reference material (BCR ® -483, extractable trace elements in sewage sludge amended soil) was used to check the accuracy of the sequential extraction procedure. For this last evaluation, only 2 metals (Cu and Zn) could have been checked, since the other metals were not certified by CRM. Good results were obtained for the first step with less than 1% relative errors and less than 20% relative errors were calculated in steps 2 and 3.

Sediment quality guidelines

The total concentrations of elements recorded in sediments were compared to French recommendations [START_REF] Geode | Suivis environnementaux des opérations de dragage et d'immersion[END_REF][START_REF] Record | Impacts écologiques de sédiments pollués extraits et déposés en milieux terrestres[END_REF] relative to dredged sediment dumping operations, which were established taken under the guidelines of the international OSPAR Convention of 1992.

Two levels (N1 and N2) were determined by the GEODE for 8 elements: As, Cd, Cr, Cu, Hg, Ni, Pb and Zn. Level N1 gives the concentrations with no or an insignificant potential impact, for which the dumping maybe authorized but a follow-up study is required from the overtaking of this threshold.

Level N2 gives the concentrations above which the dumping not maybe authorized, only if proof is brought that it is the least harmful solution for the aquatic and ground environment.

The total concentrations of elements were also compared to the sediment quality guidelines (SQGs)

for marine ecosystems, such as ERL (effects range-low) and ERM (effects range-median) values [START_REF] Long | Incidence of adverse biological effects within ranges of chemical concentrations in marine and estuarine sediments[END_REF] and TEL (threshold effect level) and PEL (probable effect level) values [START_REF] Macdonald | Development and evaluation of sediment quality guidelines for Florida coastal waters[END_REF]. These SQGs were assessed from statistical analyses of data collected in the sediments of North America and provide classifications of sediments in relation to their contamination state to prioritize concerning areas [START_REF] Macdonald | Development and evaluation of sediment quality guidelines for Florida coastal waters[END_REF][START_REF] Long | Predicting toxicity in marine sediments with numerical sediment quality guidelines[END_REF]. ERL and ERM were determined based on the chemical concentrations that are linked to adverse effects on organisms and delineate 3 concentration ranges (Byrkit, 1975[START_REF] Long | Incidence of adverse biological effects within ranges of chemical concentrations in marine and estuarine sediments[END_REF]: concentration below the ERL value representing the minimal-effects range of the sediment associated with the chemical on organisms (< 10% of biota); the concentrations equal to and above the ERL value and below the ERM value representing a possible-effects range; the concentration above the ERM value representing a probable-effects range (≥ 50% of biota). TEL and PEL represent the concentrations that are associated with the observed and non observed effects and give adverse effects that are expected to occur rarely or frequently, respectively [START_REF] Macdonald | Development and evaluation of sediment quality guidelines for Florida coastal waters[END_REF]Rahman et al., 2014a).

Sediment quality assessment

To assess the level of anthropogenic pollution of the sediments, the enrichment factors (EFi) were calculated after normalization of the total element concentrations Ci [START_REF] Salati | Assessment of heavy metal concentration in the Khoshk River water and sediment, Shiraz, Southwest Iran[END_REF][START_REF] Zahra | Enrichment and geo-accumulation of heavy metals and risk assessment of sediments of the Kurang Nallah-Feeding tributary of the Rawal Lake Reservoir[END_REF] and selection of a reference area. Normalization is defined as a procedure to correct contaminant concentrations for the influence of the natural variability in sediment composition (grain size, organic matter or mineralogy) and is recommended by the OSPAR Convention (OSPAR Commission, 2014). An abundant element of the fine fraction (rich in aluminosilicates), such as Al, Fe, Mn or Sc, is currently chosen to normalize the metal concentration (Ci) of sediment [START_REF] Zahra | Enrichment and geo-accumulation of heavy metals and risk assessment of sediments of the Kurang Nallah-Feeding tributary of the Rawal Lake Reservoir[END_REF][START_REF] Pourabadehei | Effect of the resuspension technique on distribution of the heavy metals in sediment and suspended particulate matter[END_REF][START_REF] Jafarabadi | Spatial distribution, ecological and health risk assessment of heavy metals in marine surface sediments and coastal seawaters of fringing coral reefs of the Persian Gulf, Iran[END_REF][START_REF] Gargouri | Use of sediment quality indicators for heavy metals contamination and ecological risk assessment in urbanized coastal zones[END_REF][START_REF] Jahan | Comparison of pollution indices for the assessment of heavy metals in the sediments of seaports of NSW, Australia[END_REF][START_REF] Kim | Critical evaluation of different methods to calculate the Geoaccumulation Index for environmental studies: A new approach for Baixada Santista -Southeastern Brazil[END_REF][START_REF] Savitha | Distribution of Heavy Metals in the Marine Sediments of Various Sites in Karaichalli Island, Tuticorin, Gulf of Mannar, India[END_REF][START_REF] Tunca | An ecological risk investigation of marine sediment from the northern Mediterranean coasts (Aegean Sea) using multiple methods of pollution determination[END_REF]. Here, the percentage of the 63 µm fraction (%63) that was considered to represent the fine fraction was preferred to normalize the metal concentrations. Point 1 (for harbor monitoring) and tank 1 (for laboratory experiment) were chosen as reference areas to calculate the enrichment factor as in Eq. ( 1).

EF i = (C i %63) sample ⁄ (C i %63) reference ⁄ Eq. (1)
Based on EFi values, [START_REF] Hasan | Enrichment factor and geo-accumulation index of trace metals in sediments of the ship breaking area of Sitakund Upazilla (Bhatiary-Kumira), Chittagong, Bangladesh[END_REF] suggested seven categories of sediment metal enrichment: no enrichment of sediment presenting EF < 1, minor enrichment with EF from 1 to 3, moderate enrichment with EF from 3 to 5, moderately severe enrichment with EF from 5 to 10, severe enrichment with EF from 10 to 25, very severe enrichment with EF from 25 to 50 and extremely severe enrichment with EF ≥ 50.

Ecological risk assessment

To assess the ecological risk of the sediments in relation to their metal concentrations, different indices are often used, such as individual potential ecological risk factor Er i , the toxic unit (TU) and the mean effect range-median quotient (m-ERMq). [START_REF] Hakanson | An ecological risk index for aquatic pollution control.a sedimentological approach[END_REF] was the first to develop an individual potential ecological risk factor (Er i ) by using Eq. ( 2), as follows:

Er i = Ti. Cfi Eq. ( 2
)
where Ti is the toxic response factor associated with the element i and Cfi is the contamination factor for the same element. But, as [START_REF] Brady | Development of a hybrid pollution index for heavy metals in marine and estuarine sediments[END_REF] who proposed a modified pollution index based on using the enrichment factor instead of the contamination factor Cfi to take consider the nonconservative behavior of sediments, [START_REF] Gargouri | Use of sediment quality indicators for heavy metals contamination and ecological risk assessment in urbanized coastal zones[END_REF] proposed a modified ecological risk factor.

Then, in this paper, the authors used the modified individual potential ecological risk factor mEr i , as the following Eq. ( 3):

mEr i = Ti . EFi Eq. ( 3
)
where EFi is the enrichment factor of the element i based on its normalized concentrations to percentage of fine fraction < 63 µm (Eq. ( 1)). Then, the risk factor was described by the original classification of [START_REF] Hakanson | An ecological risk index for aquatic pollution control.a sedimentological approach[END_REF] [START_REF] Hakanson | An ecological risk index for aquatic pollution control.a sedimentological approach[END_REF].

The toxic unit (TU) is related to the toxic effects of elements on living organisms and was proposed by [START_REF] Long | Predicting toxicity in marine sediments with numerical sediment quality guidelines[END_REF] as following Eq. ( 4): TU = Ci / PELi Eq. ( 4)

According to the method, TU < 0.1 is a 10% probability of being toxic, 0.11 < TU < 1.5 is a 25% probability of being toxic, 1.51 < TU < 2.3 is a 50% probability of being toxic, and TU > 2.3 is a 76%

probability of being toxic.

The mean effect range-median quotient (m-ERMq) was suggested by [START_REF] Long | Classifying probabilities of acute toxicity in marine sediments with empirically derived sediment quality guidelines[END_REF] as a method to understand the toxic effects on living organisms by taking account all the elements in the studied environment [START_REF] Tunca | An ecological risk investigation of marine sediment from the northern Mediterranean coasts (Aegean Sea) using multiple methods of pollution determination[END_REF]. This index value is determined by the following Eq. ( 5):

m-ERMq = Σ /
Eq. ( 5)

with n, the number of considered elements.

According to the method, m-ERMq < 0.1 is a 9% probability of being toxic, 0.11 < m-ERMq < 0.5 is a 21% probability of being toxic, 0.51 < m-ERMq < 1.5 is a 49% probability of being toxic, and m-ERMq > 1.50 is a 76% probability of being toxic [START_REF] Long | Classifying probabilities of acute toxicity in marine sediments with empirically derived sediment quality guidelines[END_REF].

Statistical analyses

The analyses were conducted for each experiment (annual in situ monitoring, with experiments performed by caging and in the laboratory) using the software Statview 5.0 (SAS). To assess differences in the biochemical responses between sites and between dates of sampling, data were analyzed with a one-factor analysis of variance (ANOVAs) followed by Student's Newman-Keuls post hoc tests (SNK). Pairwise comparisons were then achieved between the groups. Levels of statistical significance were set at p < 0.05.

Results and discussion

Physical chemical characterizations of sediments in both studied areas

The results of the structural parameters (TOC and fine fraction < 63 µm) indicated significantly (p < 0.001) higher levels of TOC (37.4 ± 2.6%) in harbor sediments submitted to a weaker condition of agitation (point 3) than in sediments located at points 2 (23.1 ± 6.1%) and 1 (27.5 ± 5.1%), which were under stronger agitation. No significant differences were calculated between the fine fraction levels of harbor sediments collected at point 3 (71.2 ± 20.0%), point 2 (78.6 ± 10.6%) and point 1 (80.0 ± 9.4%). In sediments monitored in all tanks of the laboratory experiment, reproducing the anode degradation, TOC and fine fraction levels were similar under all conditions of agitation, and these levels were close to those measured in harbor sediments exposed to agitation (point 2). No significant differences in the TOC and fine fraction levels were recorded during the time of both monitoring (in harbor and in laboratory).

All these results were similar to those monitored in harbor areas [START_REF] Stronkhorst | Contaminants of Concern in Dutch Marine Harbor Sediments[END_REF]. They

showed that the characterizations of sediments exposed in laboratory experiments reproduced the in situ harbor conditions well enough, with the exception of lower TOC levels recorded in the sediments of tank 3 under weak agitation.

Total concentrations in sediments

Table 1 relates the concentrations of elements measured in sediments collected in the harbor near (0 m) or distant from anodes (15 m), at point 1 (reference area), point 2 (agitated area) and point 3 (nonagitated area), during 42 months of monitoring. The first year of monitoring (Sept. 2012 to Sept.

2013) corresponding to the period prior to anode implantation, sediments were considered as homogeneous near and at 15 m away from the docks. Thus, analyses were performed only on sediments at proximity of the docks (0 m). These concentrations were compared to SQGs and GEODE levels (N1 and N2) that were only reported for Cu and Zn, since the other elements were not documented.

No significant differences in concentrations were observed in the harbor sediments of points 1 and 2.

At these points, the Cu and Zn concentrations were relatively stable during the 42 months of monitoring in sediments located at proximity of the anodes. These concentrations were 2 to 3 times lower than those measured in the sediments of point 3 (non-agitated area), in which higher significant (p < 0.0001) concentrations of Al and Fe were recorded. Then, Cu and Zn concentrations increased with time (+86% for Cu and +137% for Zn) to respectively reach 51.4 mg kg -1 and 347 mg kg -1 during the last months of monitoring. Despite its increase, Cu concentrations remained inferior to the N1 level of the French guidelines GEODE. This was not the case for Zn concentrations in the same sediments that exceeded the N1 level (276 mg kg -1 ) during the last year of monitoring (2016)(2017).

The evaluation of the sediment quality was also discussed based on the comparison of the monitored sediment concentrations with the two sets of SQGs guidelines, the ERLs/ERMs and TELs/PELs [START_REF] Long | Predicting toxicity in marine sediments with numerical sediment quality guidelines[END_REF]. Cu concentrations sometimes exceeded the TEL in the sediments of points 1, 2 and 3 and were regularly over the ERL in the sediments of point 3 during the 42 months of monitoring. The Zn concentrations that were recorded in the sediments of point 3 were over TEL, regularly close to the ERL, and during the last year (2016 -2017) of monitoring, exceeded the PEL and were close to the ERM. probable effect level, (c) effet range-low, (d)

379

Table 2 reported concentrations of elements researched in sediments exposed in the laboratory to 380 anodes in agitated (tank 2) and non-agitated conditions (tank 3). Tank 1 referred to uncontaminated 381 sediment by anodes (reference). For the harbor sediments, only the Cu and Zn concentrations could 382 have been compared to the GEODE levels, and the SQGs guidelines which were reported in the same 383

Table. 384

In the sediments that were monitored in all tanks during the 18 months of experiment, the 385 concentrations of In, Mn, Fe and Si were stable, and the Cu concentrations presented a small increase observed in the sediments of tank 1 (control) with +187% (p < 0.002), which suggested small inputs with natural seawater. In the controlled sediments (tank 1) and sediments of tank 2, no concentrations exceeded the GEODE levels and the SQGs guidelines, except the TEL of Zn for sediments collected in November 2015. Then, natural contributions of Zn (referred by the levels measured in the control tank) were considered in all tanks during the laboratory experiment. Taking account this hypothesis, sediments of tank 3 collected after 6 months of experiment presented Zn concentrations that exceeded all sets of guidelines, including ERM and the N2 level of GEODE, especially in July 2015 and March 2016. Although Zn concentrations were more important at the end of the experiment, the highest level was recorded in July 2015. The thickness of corrosion product formed on anode surface is not homogeneous and not adherent [START_REF] Deborde | Impact of Galvanic Anode Dissolution on Metal Trace Element Concentrations in Marine Waters[END_REF], which may explain the non-continuous release of elements in the water and pics of concentration in sediment before the end of monitoring.

Based on these concentrations, sediments located near the anodes in the non-agitated harbor area, may represent a possible risk of toxic effect on organisms linked to Zn concentrations (Rahman et al., 2014a;[START_REF] Tunca | An ecological risk investigation of marine sediment from the northern Mediterranean coasts (Aegean Sea) using multiple methods of pollution determination[END_REF]. These results therefore justify the following analyses conducted to assess the quality of harbor sediments.

Enrichment factors

Table 3 reported enrichment factors (EF) calculated for sediments monitored in the harbor (point 2 and point 3) and in the tanks exposed to one anode during the laboratory experiment (tank 2 and tank 3). In the harbor, EFs were also calculated for sediment located at 15 m from the anodes. The EF values helped us to identify the elements in sediments, which present an anthropogenic origin. According to the [START_REF] Hasan | Enrichment factor and geo-accumulation index of trace metals in sediments of the ship breaking area of Sitakund Upazilla (Bhatiary-Kumira), Chittagong, Bangladesh[END_REF], sediments of the harbor that were collected in proximity of the anodes in nonagitated condition (point 3) can be classified as moderate to moderately severely enriched by Cu and Zn, with respectively maximal values of EF = 10 and EF = 9 recorded at the end of monitoring. For the same period, these sediments presented minor enrichments in Al and In, which did not exceed EF of 3. At 15 m from point 3, minor to moderate Cu and Zn enrichments of sediments were calculated, which suggested widespread contamination. The results achieved for sediments that are exposed to anodes in non-agitated condition during the laboratory experiment (tank 3) showed the same levels of Zn enrichment factors compared to in situ enrichment. The highest Zn EF were measured earlier, in July (EF = 11) and September 2015 (EF = 6). The sediments that were collected during the months following this operation were less impacted and presented moderate enrichment in Zn (EF values of 3 to 5), which suggested a dispersion of Zn in seawaters. However, if minor Al and In enrichment factors (EF inferior of 4) were observed in the same sediments that were exposed to anodes in laboratory conditions, these sediments were not or were slightly affected by a Cu enrichment (EF < 2).

In their study, [START_REF] Gabelle | The impact of aluminium sacrificial anodes on the marine environment: A case study[END_REF] related minor to moderate Al enrichment factors of harbor sediments collected near or at 20 m from anodes with a maximal average value given for the Al EF of 2.8 ± 0.2. During our work, the same Al enrichment factors were recorded in sediments that were monitored either in a harbor environment or under laboratory exposition conditions. Al is the major component of the anode with 93.2%, and because of its mainly non soluble form, we might think that the Al dissolved from galvanic anodes would have been induced quickly to accumulate into sediments.

Indeed, Al is known for its colloidal behavior, which conducts it to form rapid aggregates during the flocculation process. Al is integrated to sediment when the forming precipitates (called flocs) settled according to their weight. However, Al enrichment factors recorded in all sediments of both experiments (in situ and in laboratory), remained inferior to 3 and were considered to be minor. One explanation could be linked to the low kinetic of aggregation, which would conduct to the continuous dispersion of Al in the water even under weak agitation of the middle. The highest levels of metallic enrichment in harbor sediments were associated with Zn and Cu, while these elements are, respectively, the second (with 2.5 to 6.5%) and one of the last elements (with 0.006%) composing the anode. Such levels of enrichment were already given for harbor sediments but were often associated with other sources of contamination, such as antifouling paints [START_REF] Singh | Leaching of copper and zinc from spent antifouling paint particles[END_REF][START_REF] Pourabadehei | Effect of the resuspension technique on distribution of the heavy metals in sediment and suspended particulate matter[END_REF][START_REF] References Amara | Antifouling processes and toxicity effects of antifouling paints on marine environment. A review[END_REF] or urban runoff [START_REF] Macdonald | Development and evaluation of sediment quality guidelines for Florida coastal waters[END_REF][START_REF] Nabizadeh | Study of heavy metals in urban runoff[END_REF][START_REF] Reddy | Removal of heavy metals from urban stormwater runoff using different filter materials[END_REF]. If the results of our laboratory experiment confirmed the same enrichment of Zn in sediments exposed to the anode (without ruling out the hypothesis of another source of contamination), they clearly did not attribute the Cu enrichment of sediment to the anode dissolution.

The high levels of Cu recorded in the sediments harbor probably is associated with the other sources of contamination linked to the harbor environment cited previously. Our attention was retained by the

In results. This element represents 0.01% of the anode composition but minor enrichments of harbor sediments were also observed, and it was confirmed or enhanced during the simulation of anode dissolution performed during our laboratory experiment. This element is not usually met in the harbor environment, which was confirmed at the beginning of our monitoring by very low values of EF. The origin of this element in the marine environment may be associated to e-waste generated by electronic manufacturing [START_REF] Chen | Fractionation and release behaviors of metals (In, Mo, Sr) from industrial sludge[END_REF][START_REF] Jensen | The mobility and plant uptake of gallium and indium, two emerging contaminants associated with electronic waste and other sources[END_REF]. In our study, we demonstrated that this element might be considered to be a tracer of the anode dissolution, especially for sediments that are monitored in non-agitated environments.

Ecological risk assessment

To assess the effects of elements on the living organisms in the different sediments, we selected three indicators that are usually used to assess the effect of sediment pollution on the ecosystem: the modified ecological risk assessment (ER i ), the toxic unit (TU) and the mean effect range-median quotient (m-ERMq) (Table 4). The results associated with sediments collected in the harbor and during the laboratory experiment were reported in Table 4, but only for Zn and Cu since no references were given for the other elements constituting the anode, such as Al and In. In this study, sediments exposed to the condition of agitation (point 2 and tank 2) presented low ecological risk (mEi < 40).

The same observation was given for sediments under non-agitation conditions, except for one sediment collected at the end of the monitoring period (February 2017), which presented a moderate risk with mEi of 49. The sediments in majority showed a TU < 1.5 except for those exposed to the anode after July 2015 for which the probability of being toxic increased to 50% with a maximum of 76%. The results of m-ERMq indicated the global toxic effect of Cu and Zn. The majority of sediments (whatever the condition of agitation) were classified on the second grade of the 4-level scale with a 21% toxicity level. However, this level of toxicity increased to 49% for sediments exposed to non-agitated conditions after one year of anode implementation in the harbor or after 9 months (July 2015) for the laboratory experiment. These results confirmed that the condition of agitation considerately influenced the potential effect of sediments to organisms, especially when these sediments were collected near the anodes. For Al and In, which were found to uncontaminate to moderately contaminate the sediment, it is difficult to assess their possible effect on organisms on the basis of the previous indicators. Previous works demonstrated the biological effects of Al in the aquatic environment but only in low pH conditions of the environment (pH < 4.5 where Al 3+ form of Al is dominant), which was not the environmental condition of these works [START_REF] Gostomski | The toxicity of aluminum to aquatic species in the US[END_REF][START_REF] Rosseland | Environmental effects of aluminium[END_REF][START_REF] Wilson | Chapter 2: Aluminum[END_REF]. 24

Results of sequential extractions

Although the SQGs guidelines and the previous quality indices helped us to identify the elements and sites linked to a potential risk to organisms, they did not identify the chemical form of the element that may be responsible for causing toxicity. Therefore, a study on the speciation of each element that was released from the anode and eventually transferred to the sediments was conducted during both studies (in the harbor and in the laboratory). Then, sequential extractions were performed to assess the bioavailability of the elements by exploring their concentration in four fractions of the sediments:

acid-soluble fraction (F1), reducible fraction (F2), oxidizable fraction (F3) and residual fraction (F4). In harbor sediments, Al, Cu, Fe and Si were principally present in the F4 (residual, 78 to 98%) fraction, Mn was mostly scavenged in the F1 (acido-soluble, 45%) fraction and Zn was distributed between the F4 (residual, 60%) fraction and equally in fractions F1, F2 and F3 (10 to 17%). After 42 months of monitoring, Mn and Si distributions remained stable in all sediments. During the same period, the oxidizable fraction (F3) of sediment collected at point 3 (under non-agitation conditions) was enriched by Al, Cu and Zn while Fe was decreased in this fraction. Regarding the results of the extractions performed in sediments exposed to anodes for 18 months in tank 3 (which reproduced the harbor conditions of point 3), Al and Zn were similarly enriched by the F3 fraction and, at the same time, a decrease in Fe occurred. The previous results showed higher organic matter, minor Al and moderately severe Zn enrichment of sediments located at the proximity of the anode under nonagitated conditions (point 3 and tank 3). We could suppose that Al was mobilized with time in an oxidizable fraction of the sediment when organic matter increased. Humic substances dissolved in seawater may represent important ligands for trace metals and stability of the complexes as functions of the metal, such as the following decreasing scale proposed by [START_REF] Yang | Metal Complexation by Humic Substances in Seawater[END_REF]: Cu > Zn > Fe > Al. The authors demonstrated the existence of competition between the metals for the same binding sites of humic substances and a decrease of Fe linked to humic substances with the addition of other metals. Mobilized in organic matter and the sulfides fraction (F3) of harbor sediments, Al, Cu and Zn, could be released under oxidizing conditions, which can occur if the sediment is resuspended, such as by dredging, currents, etc. [START_REF] Morillo | Heavy metal distribution in marine sediments from the southwest coast of Spain[END_REF]. In point 2 of the harbor, the evolution of the element distribution was different than in point 3, since enrichments of the F2 fraction by Zn, Fe and Cu as well as an enrichment of the F1 fraction by Zn were observed. Fraction F2 is representative of the potential element mobility under changing conditions. These differences of element mobilization were certainly linked to the difference of agitation conditions between point 3 and point 2. Actually, in the area of point 2, sediments are submitted to a resuspension technique two times per year. [START_REF] Pourabadehei | Effect of the resuspension technique on distribution of the heavy metals in sediment and suspended particulate matter[END_REF] showed that the effect of the resuspension technique reduced the concentration of metals in sediment and suspended matter by 24%. However, these enrichments were not significantly reproduced in tank 2 during the laboratory experiment, although, we noticed that Zn in control tank (tank 1) tended to enrich the F1 fraction with time instead of the F3 fraction as in tanks 2 and 3 exposed to the anode. This could suggest different sources of Zn in tank 1 (Zn associated with a natural origin) and Zn distribution in the sediments of tanks 2 and 3 (which would be linked to anode dissolution). Enrichments of Al (and In) did not present an ecological risk to the marine environment, while this element is the principal component of the anode (> 96%). The results obtained in our present study confirmed the hypothesis suggested by [START_REF] Deborde | Impact of Galvanic Anode Dissolution on Metal Trace Element Concentrations in Marine Waters[END_REF] that the corrosion product induced continuous metal diffusion towards the water column and the sediment. They added that once released, metals exhibit different behaviors related to their interaction with the environment (difference of chemical forms). Over time, the distribution study of the metals into the different fractions of sediments showed that released Zn and Al were transferred to the oxidizable fraction. These results supposed a potential mobilization of these elements from the sediment if physical-chemical changes would occur (for example during dredging operations).

On the other hand, these results suggested that In could be a good tracer of anode dissolution in harbor sediment monitored in non-agitated areas, under long-term monitoring condition (more than one year).

All these results needed to be completed by further and sediment bioassays performed on living organisms to assess the potential toxic effects of Zn, Al and In that were released from the galvanic anode and that accumulated in sediments. They will be followed by the results obtained on organisms (mussels and sponges) that were exposed during the same experiments.

Fig. 1 :

 1 Fig. 1: Location of the sampling points monitored in port of Calais.

Fig. 2 :

 2 Fig. 2: Experimental devices with or without a galvanic anode to reproduce anode degradation under laboratory conditions.

  : < 40, 40 to 80, 80 to 160, 160 to 320 and ≥ 320, which involved low potential ecological risk, moderate potential ecological risk, considerable potential ecological risk, high potential ecological risk and very high potential ecological risk. In this work, only Cu and Zn could have been associated with Ti, since these factors were only established for the following metals: As = 10, Cd = 30, Cr = 2, Cu = Pb = 5, Hg = 40 and Zn = 1

Fig. 3

 3 Fig.3shows the concentrations of elements (Al, Cu, Fe, Mn, Si and Zn) that were extracted from the F1, F2, F3 and F4 fractions of sediments exposed to anodes in agitated and non-agitated conditions during harbor monitoring (left graph) and the laboratory experiment (right graph). The sequential extraction of In was not conclusive due to the low concentrations detected in each fraction (inferior to the detection limits).

Fig. 3 :

 3 Fig. 3: Concentrations of elements constituting galvanic anodes researched in different fractions of sediments

Table 1 :

 1

	Concentrations are compared to the national N1/N2 guidelines (GEODE levels) and SQGs.

Concentrations of elements (constituting galvanic anodes) researched in sediments collected in situ harbor monitoring in the reference area (point 1), near anodes (0 m) or distant from anodes (15 m) under agitated (point 2) and non-agitated (point 3) conditions of the harbor between September 2012 and September 2013 (before anode implementation) and between January 2015 and June 2017 (after anode implementation).

Table 2 :

 2 Concentrations of elements constituting galvanic anodes researched in sediments exposed to anodes 406 under agitated and non-agitated conditions during a laboratory experiment. Concentrations are compared to the

	407
	408	national N1/N2 guidelines (GEODE levels) and SQGs.

(p < 0,008). At the same time, concentrations of Al increased in tank 3 (p < 0.0001) and in tank 2 (p <

0.001), and the higher rises were given for Zn concentrations in the sediments of tank 3, with +1130%

(p < 0.0001), and tank 2 with +337% (p < 0.0005). A few increases of Zn were simultaneously

Acknowledgments:

The authors would like to thank the Société d'Exploitation des Ports du Détroit (Port of Calais), especially Patrick Honoré for his participation in the project management; the

0.8 0.9 0.9 0.7 0.8 0.9 0.8 -0.9 1.0 1.1 1.3 0.7 0.7 Jul-16 0.8 0.9 0.8 1.0 0.8 0.9 0.7 0.9 0.8 0.9 1.2 1.2 0.9 1.0 Feb-17 1.0 0.6 1.1 0.9 1.0 0.5 0.9 0.6 0.9 0.6 1.2 2.0 0.9 0.5 Jun-17 0.8 0.7 1.4 0.2 0.8 0.5 0.9 -0.9 0.6 Sep-14

Non-agitated condition (point 3)

LABORATORY EXPERIMENT

Agitated condition (tank 2)

In Mn Si Zn