Cathodic protection by zinc sacrificial anodes: Impact on marine sediment metallic contamination

Journal of Hazardous Materials 167 (2009) 953–958

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Cathodic protection by zinc sacrificial anodes: Impact on marine sediment metallic contamination C. Rousseau ∗ , F. Baraud, L. Leleyter, O. Gil Equipe de Recherche en Physico-Chimie et Biotechnologies (ERPCB) EA3914, Université de Caen-Basse Normandie – Bd du Maréchal Juin – Bât. Sciences 2, 14000 Caen, France

a r t i c l e

i n f o

Article history: Received 22 September 2008 Received in revised form 21 January 2009 Accepted 21 January 2009 Available online 30 January 2009 Keywords: Sediment Seawater Zinc Chemical extraction Sacrificial anode

a b s t r a c t Cathodic protection by sacrificial zinc anodes is often applied to prevent immerged metallic structures from corrosion. But this technique induces the zinc anodes dissolution, which can induce marine sediments and seawater contamination. A large scale experiment, in natural seawater, was conducted during 12 months, in order to evaluate the potential environmental impact of this continuous zinc dissolution, and of some necessary cleaning operations of the anodes surfaces. The heavy metal (Cr, Cu, Pb and Zn) concentration in water and sediment samples was monitored. A sequential extraction procedure was applied on sediment samples to differentiate the zinc mobile fractions from the residual one. A significant increase of zinc concentration was observed in water as well as in the surface sediments under the specific operating conditions. Sediments then become a secondary pollution source, as the sorbed labile zinc can be remobilized to seawater. © 2009 Elsevier B.V. All rights reserved.

1. Introduction A successful way to prevent corrosion of metallic structures in seawater is to apply cathodic protection. The technique consists in lowering the structure electrical potential into its protection range, which can be realised either by galvanic coupling with a sacrificial anode (i.e., zinc or aluminium) or by the application of a fixed current density with an inert anode. The cathodic currents induce the reduction of dissolved oxygen (1), generating hydroxyl ions at the very near polarized surface. This reaction then increases the interfacial pH and induces the precipitation of an inorganic layer, mainly constituted of calcium carbonate and some magnesium components (2 and 3). This mixed deposit is generally called “calcareous deposit”. As this non-conductive deposit progressively covers the surface, the overall rate of the cathodic reaction (oxygen reduction) decreases. The anodes dissolution (i.e., zinc oxidation (4)) rate is then lowered, when galvanic coupling is used to apply cathodic protection. O2 + 2H2 O + 4e− → 4OH−

(1)

Mg2+ + 2OH− → Mg(OH)2 (s)

(2)

Ca2+ + CO3 2− → CaCO3 (s)

(3)

Zn → Zn

2+

+ 2e

−

(4)

∗ Corresponding author. Tel.: +33 2 31 56 74 91; fax: +33 2 31 56 73 03. E-mail addresses: [email protected], rousseau [email protected] (C. Rousseau). 0304-3894/$ – see front matter © 2009 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2009.01.083

Thus, the anode dissolution rate varies according to the cathodic protection steps. At the very beginning, as the current demand is the most important, the dissolution rate is the highest. Then, when the calcareous deposit is forming and progressively covering the entire surface, the current demand decreases, which then reduces the anode dissolution. Moreover, other phenomena can occur at the anodes surfaces. For instance, suspended matter can be deposited, or a layer of zinc oxides can be formed at the anodes surfaces [1], isolating the sacrificial anode from the electrolyte medium. Such phenomena reduce the cathodic protection efficiency (as the potential then increases) but also consequently reduce the anode oxidation and consequently, the zinc release. Under controlled experimental conditions it might be necessary to operate some cleaning of the anodes surfaces, in order to remove these layers and to maintain an efficient cathodic potential. In harbours, these cleanings are naturally realised by the tide and the current which tear out the scale. The zinc oxides, then liberated, can be dissolved, which then induces an important input of zinc into water. The zinc sacrificial anodes are widely used to protect ships and harbour structures, which raises the question of local water contamination by the solubilized zinc. In harbours, the increasing numbers of ships using zinc sacrificial anodes, combined to the lower renewal of the seawater by tide, contributes to an increase of the water zinc level. Concentrations up to 25 ␮g L−1 , depending of the tide exposition, have been observed by Bird et al. [2], whereas the zinc average concentrations in the open ocean are inferior to 0.5 ␮g L−1 [3]. This input may cause ecological damage, as zinc is recognized to be toxic up to 5 ␮g L−1 by the OSPAR

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Fig. 1. Scheme of the experimental tank (a) position of the immerged samples, (b) position of the buried samples

Convention ecotoxicological assessment criteria (EAC) [4]. Moreover, zinc is bioaccumulable and can contaminate the food chain [5]. Some of the solubilized zinc can be transferred to the sediment phase, via sorption phenomena onto solid particles. In harbours, zinc concentration in the sediment phase can be 2.5 times higher than in non-contaminated areas [2]. As other heavy metals, the zinc toxicity depends on its total concentration but also on its mobility and reactivity with the ecosystem constituents. Therefore, adsorbed metals onto solid particles are potentially available as they may be dissolved due to changes in the physicochemical properties of the environment, like salinity, pH, redox conditions and/or organic chelators occurrence [6,7]. This potential mobilization (or lability) can be estimated by chemical extraction procedures. Thus, sequential extraction procedure (several steps) are used to differentiate mobile from residual fractions, and to characterize the different labile fractions [6,8–10] as: exchangeable fraction (weakly bound elements in equilibrium with the dissolved phase), acido-soluble fraction (very sensitive to pH variations), reducible or oxidable fractions (very sensitive to redox conditions and microbiological activity). Sequential extractions are then useful for solid speciation of particulate elements, to study the origin, the fate, the biological and physicochemical availability and transport of sorbed elements. According to several studies on metal fractionation, zinc is one of the most mobile metals as mainly associated with the non residual fraction sediments [11–16]. Generally, zinc is mainly associated to the acido-soluble and reducible fractions [2,11,16,17]. Exchange reaction between the sediment and water phase can then occurred as these labile fractions can liberate elements under changing physicochemical conditions (pH, redox potential variation), as during harbour dredging operations for instance. Then, contaminated sediments can be considered as a secondary pollution source towards water. In order to evaluate the potential zinc contamination of sediments and seawater, due to the use of zinc sacrificial anodes, a large scale experiment, with natural seawater, was conducted during 12 months. Cathodic protection was applied on carbon steel structures thanks zinc sacrificial anodes. This experiment was designed to study the cathodic protection and the calcareous deposit formation on carbon steel samples immersed in natural seawater and buried in sediment. One anode per designed steel structure was used to maintain the potential at −1 V/AgAgCl, which is the value commonly used to maintain the cathodic protection. Theses structures were designed to allow enough sampling (and deposit analysis) during the whole experiment. The mass of anodes to use is then depending of the metallic surface to protect. A sequential extraction procedure was applied on sediment samples to evaluate the zinc mineralogical fractionation. Some other heavy metals (Cu, Cr, Pb and Zn) concentrations, in water and sediment samples, were also monitored to detect a possible influence of the cathodic protection on their distribution.

2. Experimental 2.1. Cathodic protection with zinc sacrificial anode A large scale experiment was conducted, during 12 months, in a tank (about 9 m3 ), supplied with decanted natural seawater, which was pumped at 400 m from the shore (stable pH, equal to 8.1) at the marine station of Luc-sur-Mer (Normandy, France). The water temperature evolves with seasons from 14 ◦ C to 20 ◦ C. In the experimental basin, a continuous seawater circulation is maintained at 5 m3 h−1 which allows a good oxygenation of the water, but let the suspended matter settle down. An additional seawater inlet (freshly pumped water), which represents 5% of the seawater rate, leads to a total tank renewal each 4 h. A 70 L container, filled with natural marine sediments (obtained from the seawater settling, particles diameters <60 ␮m), is immersed in the tank. Some of the steel samples are buried in the sediments and the others are immersed in water (Fig. 1). All these samples are corrosion-protected thanks to 10 sacrificial zinc anodes (one anode per structure). The zinc anodes are provided by Bac Corrosion. All the impurities (Al, Cd, Fe, Pb . . .) concentrations are less than 0.1% of the anode. The resulting potential value is −1 V/AgAgCl, which is the common value used for cathodic protection. In order to keep the potential at this value, four manual cleaning operations of the anodes were necessary and realised as soon as the potential was over than −0.9 V/AgAgCl (after 100, 200, 275 and 350 days). This loss of protection is due to an excess of zinc oxides on the anodes [1]. The quantity of anodes decreases during the experiment as two anodes (along with two carbon steel structures) are removed at each sampling operation, after 1, 3, 6, 9 and 12 months of cathodic protection [18]. This freTable 1 Limit of detection and analyse of reference material LGC 6019 (water). Element

Limit of detection (␮g L−1 )

Certified values (␮g L−1 )

Cr Cu Pb Zn

6 2 12 14

0.78 15.4 5.2 59.7

*

± ± ± ±

0.2 1.5 0.3 2.5

Measured values (␮g L−1 ) *

14 ± 1.3 *

63 ± 1.0

Under the limit of detection.

Table 2 Limit of detection and analyse of reference material HR1 (sediment). Element

Limit of detection (␮g g−1 )

Certified values (␮g g−1 )

Cr Cu Pb Zn

0.8 2.8 2 2.5

126 79.9 139 1105

± ± ± ±

45 11.4 37 173

Measured values (␮g g−1 ) 124 67.4 130 1028

± ± ± ±

1.6 1.6 3.8 13.9

C. Rousseau et al. / Journal of Hazardous Materials 167 (2009) 953–958

quency is justified by the calcareous deposit evolution on the steel surfaces that is monitored at the same time [18]. 2.2. Seawater and sediments sampling and characterization Seawater (SW) and sea sediments samples (SS) were periodically collected and analysed from the very beginning of the experiment (SW0 and SS0), then after 1 (SW1 and SS1), 3 (SW3 and SS3), 6 (SW6 and SS6), 9 (SW9 and SS9) and 12 (SW12 and SS12) months of cathodic protection. Sediments were collected in three zones in the container: “superficial” sediments (in contact with water), sediments closed to “steel” (15 cm from the water/sediment interface) and “deep” sediments (30 cm from the water/sediment interface). Additional samplings were also realised after the last anodes cleaning operation (350 days), when only two anodes remained immersed. Water and superficial sediments were then collected after 1, 2, 3, 4 and 11 days after this cleaning operation. After collection, sediments were air-dried at 30 ◦ C and stored at 4 ◦ C in polypropylene bottles. Cu, Cr, Pd and Zn concentrations were determined after microwave assisted acid digestion (HNO3 and HCl), by inductively coupled plasma atomic emission spectrometry (ICP-AES) (Varian Vista-MPX). These elements were also quantified in seawater by ICP-AES analyses. The quality control of ICP-AES data was assessed by the analysis of blank reagents and calibration standards prepared with commercially available solutions (Varian solution standards). Accuracy of ICP-AES measurements was determined using various certified reference materials (HR1 for sediment analyses and LGC 6019 for water). These data and the limits of detection are presented in Tables 1 and 2. Other sediments characterizations were also performed. The organic matter percentage was determined by loss-on-ignition at 375 ◦ C during 16 h [19]. The calcium carbonate rate was evaluated by the CO2 volume free method after acid dissolution. The nature of clays was characterized by X-ray diffraction (Philips, PW 3040/00 X’Pert MPD/MRD). Sediments were also leached by an optimized sequential chemical extraction procedure [6,7]. This procedure selectively and efficiently dissolves all the chemical constituents of the sediments, in the following order: F1: water soluble (released by ultrapure water), F2: exchangeable (extracted with a magnesium nitrate solution), F3: acido-soluble (leached by an acid/acetate buffer), F4: reducible (extracted with hydroxyl ammonium, oxalic acid and ascorbic acid), F5: oxidable fractions (released by hydrogen peroxide-ammonium acetate extraction) and F6: sum of all labile 5 fractions ( i=1 Fi ). 3. Results 3.1. Heavy metals (Cu, Cr, Pb and Zn) concentrations in seawater In Table 3 are presented the copper, chromium, lead and zinc concentrations (mg L−1 ) in the various seawater samples. The average values observed by Brown et al. [3] in open ocean are also reported. Table 3 Cr, Cu, Pb and Zn concentration (mg L−1 ) analysed in the seawater samples and average values in open ocean according to Brown et al. [3].

SW0 SW1 SW3 SW6 SW9 SW12 Open ocean

Cr

Cu

Pb

Zn

0.04 0.06 0.05 0.05 0.04 0.09

0.00 0.00 0.00 0.00 0.00 0.00

0.07 0.06 0.04 0.08 0.04 0.13

0.04 1.09 0.48 0.52 0.71 0.24

3 ×10−4

1 ×10−4

2 ×10−6

5 ×10−4

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For all the reported elements, the levels were superior to the values proposed by Brown et al. for open ocean [3]. This was not surprising as the water used in this experiment was sampled near the shore and is probably contaminated by various human activities. Along the experiment, no significant concentration variation was observed for any of the elements analysed, except for zinc, which presented an important increase during the first months of the experiment. The zinc initial concentration was 0.04 mg L−1 . This level is high if compared to the average values generally observed in open ocean (Table 3) but is not unusual in harbour. Indeed, Bird et al. reported values up to 0.03 mg L−1 in south coast of England harbours [2]. After 1 month of cathodic protection, an important increase of zinc concentration was noticed as the concentration value (SW1: 1.09 mg L−1 ) was about 30 times more than the initial one (SW0: 0.04 mg L−1 ). This increase was undoubtedly correlated to dissolution of the 10 zinc anodes used to apply the cathodic protection (reaction (4)). Indeed, the first month of experiment corresponded to the cathodic protection establishment, when the current demand, and so the anode dissolution, are both at their higher rate. These results differ from those observed by Leleyter et al. [16], which presented no significant increase of the water zinc concentration after 1 month of cathodic protection, despite, the dissolution of zinc anodes is recognized to contribute to the input of metals to coastal waters, as shown by Bird et al. [2], Wagner et al. [20] or inside the ballast tanks shown by Jelmert and Van Leeuwen [21]. Leleyter et al. [16] values might result from an initial zinc water contamination which could dissimulate the impact of the sacrificial anodes dissolution on the seawater. In the present experiment, this important zinc contamination is only observed during the first month. Indeed, a regular decrease of the water zinc concentration was observed (SW3: 0.48 mg L−1 to SW12: 0.24 mg L−1 ) during the following months (Table 3), except for the 9 months sampling, for which a low increase, compared to the previous sampling, is detected (SW9: 0.71 mg L−1 ). Various hypotheses could explain these results. First, after 3 months, then 6 months, the current demand is reduced of about 90% [18]. This decrease results from the calcareous deposit formation on the protected structure surface, which limits the cathodic (on structures) and anodic (on anodes) reactions: the anodes dissolution is then reduced too, which limits the zinc input into water. Moreover, at each collection time, two anodes were removed from the tank [18]. Then, after 3 months and 6 months, respectively only eight and six anodes remained immersed, which contributes to lower the release of zinc into water. The water renewal also limits the increase of water zinc concentration by a dilution effect. Furthermore, as previously described, exchange reaction between water and solid phases (suspended matters, sediments) can occur. For instance, at pH 8.2, when Zn2+ concentration exceeds 1.16 mg L−1 , it can precipitate as Zn(OH)2 [21], which then influences the apparent variation of water zinc concentration. A cleaning operation was performed on the anodes 5 days before the 9 months sampling (t: 275 days). During this operation, the zinc oxides layer is removed from the anodes surface, but cannot be removed from the tank and is then dispersed into the experimental medium. This obviously induces an important increase of zinc level in the liquid and solid phases, as this layer is probably partly dissolved in the water. This explains the high value measured in water 5 days after this cleaning (SW9: 0.71 mg L−1 ). At the end of the experiment, the water contamination by zinc coming from anode dissolution is actually significant as, after 12 months, the zinc concentration observed (SW12: 0.24 mg L−1 ) is more than six times higher than the initial value (SW0: 0.04 mg L−1 ), despite partial water renewal and various phenomena that tend to lower water zinc levels.

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Table 4 Zinc concentration in water (mg L−1 ) and in superficial sediments (␮g g−1 ) after cleaning of the two remaining anodes. Delay following cleaning step (day)

Zinc concentration in water (mg L−1 )

Zinc concentration in superficial sediments (␮g g−1 )

0 1 2 3 4 11

0.33 0.80 1.02 1.03 1.20 0.24

830 1490 2700 2670 1780 500

The impact of the anode cleaning operation was also evaluated at the end of the experiment (350 days) before the last water and sediment collection (SW12 and SS12), when only two anodes were still immersed. The results are presented in Table 4. During the first 4 days following the cleaning operation, the zinc concentration progressively increases in water (from 0.33 mg L−1 to 1.20 mg L−1 ). This variation can be due to the progressive dissolution of some zinc solid particles that covered the anode surface (i.e., zinc oxides layer, [1]). However, 11 days after the cleaning operation, the zinc concentration is back to its initial value (0.33 mg L−1 ), suggesting that the anode cleaning step only induces a very punctual water contamination, by enhancing the zinc solid compounds dissolution. This contamination is rapidly attenuated by the dilution effect and/or zinc sorption phenomenon onto sediments. It is important to note that all the zinc concentrations observed in this experiment (from SW0 to SW12) exceed the OSPAR Convention provisional ecotoxicological assessment criteria for seawater (5 × 10−3 mg L−1 ) [4]. Moreover, we can conclude that the sacrificial anodes used in our experimental conditions induce a significant increase of the water zinc concentration. 3.2. Heavy metals (Cu, Cr, Pb and Zn) concentration in sediment 33% of calcium carbonate and 4% of organic matter were measured in the dried sediments used in the experiment [18]. The high percentage of calcium carbonate is due to the erosion of marine shells collected with sediments. The X-ray diffraction analyses indicate the presence of different clays: kaolinite, illite, chlorite, and interstratified layers of illite and chlorite. These detected clays are generally present in marine sediments of the pelagic zone [3]. Heavy metals (Cu, Cr, Pb and Zn) concentrations in sediments are presented in Table 5. Copper and lead contents determined in Table 5 Cr, Cu, Pb and Zn concentration (␮g g−1 ) analysed in the sediments samples and average values in the North Atlantic continental seabed according to Stevenson [22].

SS0 superficial SS0 deep SS1 superficial SS1 steel SS1 deep SS3 superficial SS3 steel SS3 deep SS6 superficial SS6 steel SS6 deep SS9 superficial SS9 steel SS9 deep SS12 superficial SS12 steel SS12 deep Stevenson [22]

Cr

Cu

Pb

Zn

36 35 39 40 41 47 41 46 42 46 46 46 45 53 43 49 47 55

26 24 26 27 25 20 21 21 18 18 19 18 21 19 18 19 20 15

37 37 38 40 38 41 34 35 36 38 35 38 38 40 34 35 38 11

138 126 757 135 142 537 106 115 153 124 127 750 142 133 504 129 133 57

Fig. 2. Total zinc concentration (␮g g−1 ) in sediments, and zinc concentration in seawater values (mg L−1 ) into circles.

these sediments were higher than the baseline values proposed by Stevenson for North Atlantic non-polluted sediments [22]. These sediments can then be considered as moderately polluted by Cu and Pb. Their concentrations remain stable all along the experiment and at the different depth observed. The cathodic protection had no influence on these elements level onto the sediment. The zinc total concentrations in sediments are presented on Fig. 2, in order to visualise zinc spatial and/or temporal evolution. Variation of zinc content was only observed in the surface sediment. The concentration measured in the “steel” samples (near structures) and “deep” samples (30 cm from the interface) show no significant variation, then the discussion will focus on results obtained for the superficial sediments. Variations of sediments zinc content versus time are noticed, following more or less the same pattern observed for water (water values presented into circles on Fig. 2). An important level increase is observed after the first month (from 138 ␮g g−1 (SS0) to 757 ␮g g−1 (SS1), Table 5 and Fig. 2). This can probably be explained by some zinc sorption phenomena onto the surface sediment or by the formation of zinc solid compounds (which settled down). This can be correlated to the important water zinc content increase previously described. During the 5 following months, the zinc level progressively decreases (from SS3: 537 ␮g g−1 to SS6: 153 ␮g g−1 ), and is back to its initial value after 6 months. This can only be explained by a desorption and/or redissolution process of zinc associated to the sediment, towards the water phase, as deeper sediment content remains unchanged. Moreover, such a phenomenon could contribute to maintain or slightly increase zinc water content from SW3 to SW6, as it has been noticed (Table 3, Fig. 2). This suggests that the sediments can buffer the pollution: when the water is contaminated, the sorption phenomenon reduced the water levels. But then, the phenomenon being reversible, sediments become a pollution source that can increase water metal level. After 9 months, a very important increase of zinc level is observed in the superficial sediments (SS9: 750 ␮g g−1 ) compared to the previous sampling (SS6: 153 ␮g g−1 ). As for water, this value might probably be attributed to the anode cleaning operation, realised 5 days before. As previously noticed, zinc solid compounds are probably released during this operation, being partly dissolved and/or settled down. This additional zinc is then co-analysed with the sediment and cannot be discriminated from the zinc actually sorbed onto the sediment phase. Thus, superficial sediment samples were collected (as for water) at the end of the experiment to better evaluate the anode cleaning impact. Zinc level increases during the first 2 days following

C. Rousseau et al. / Journal of Hazardous Materials 167 (2009) 953–958

Fig. 3. Quantity of zinc, in ␮g g−1 , leached from the different mineralogical fractions of the superficial sediments (residual = total − F6).

the cleaning (from 830 ␮g g−1 to 2700 ␮g g−1 , Table 2). However, 11 days after the cleaning operation, the zinc concentration (500 ␮g g−1 ) is back (or inferior) to the value determined before the cleaning (830 ␮g g−1 ). The zinc solid compounds or the sorbed zinc appear to be rapidly solubilized or desorbed into water. Finally, anode cleaning operation has a limited influence on the sediment contamination. At the end of the experiment, the zinc concentration (SS12: 500 ␮g g−1 ) is about three times higher compared to its initial value (SS0: 138 ␮g g−1 ), in spite of a limited anode dissolution (cathodic protection being stable) and a limited number of remaining anodes (two at this moment). In our experimental conditions, the sacrificial anodes induce a significant increase of the sediment zinc content. These results are in agreement with other studies on zinc contamination by sacrificial anodes [2,16], which also revealed a superficial sediment contamination. In fact, Bird et al. realised a depth profile of zinc in sediment, in a semi-enclosed marina on the south coast of England [2]. This zinc profile presents a sub-surface maximum zinc content (225 ␮g g−1 ) at the water/sediment interface and less than 150 ␮g g−1 at 20 cm from the interface. 3.3. Zinc partitioning in superficial sediments To estimate heavy metal mobility and bioavailability in sediments, the superficial sediments samples were leached by an optimized sequential chemical extraction procedure [6,7]. As for their total concentrations in sediments, no variation was observed for Cu, Cr and Pb fractionation scheme all along the experiment and in the three zones of the container [18]. The cathodic protection had no influence on the mineralogical fractionation of these elements. The situation is quite different for zinc, for which significant variation was detected in the superficial sediments (whereas zinc distribution remained stable in “deep” and “steel” samples). Then, the discussion will focus on the superficial sediments results. The zinc quantity (␮g g−1 of dry sediment) leached from the different mineralogical fractions is represented in Fig. 3; percentages values ( = ratio with the total concentration) are also calculated and represented in Fig. 4. Zinc is initially mainly scavenged in two fractions: the most important part is present in the reducible fraction (F4), with 60 ␮g g−1 (42%) and in the acido-soluble fraction (F3), with 30 ␮g g−1 (23%). This kind of distribution is not unusual for zinc and has been reported in various studies [2,11,15–17]. After 1 month, zinc is mainly associated to the acido-soluble (F3) and reducible (F4) fractions. The zinc quantity in these two fractions has increased (F3: 310 ␮g g−1 ; F4: 228 ␮g g−1 ) as a consequence of

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Fig. 4. Distribution of zinc, in %, in the different mineralogical fractions of the superficial sediments (residual = total − F6).

the total zinc content increase, as previously noticed (Fig. 2; Section 3.2). But, contrary to the initial fractionation and according to Leleyter et al. [16], the dominant fraction is the acido-soluble one (41%), followed by the reducible (30%) fraction. It is important to note that, after 1 month of experiment, no zinc was leached by water (water soluble fraction F1: 0%). This suggests that the additional zinc-associated sediment (and co-analysed), resulting from contamination by the anode dissolution, is either actually sorbed onto the sediment and/or result from the settle down of some zinc solid compounds, formed on the anode surface layer, that are not water soluble. After 3 and 6 months, the zinc quantity leached by the sequential extraction (F6) decreases, like the total zinc content in sediment. However, after 3 months, the zinc quantity decreases in F3 (from 310 ␮g g−1 to 230 ␮g g−1 ) whereas it stays stable in F4 (about 220 ␮g g−1 ), suggesting that the decrease of the total zinc content is due to the leach by the seawater of the most labile zinc (fraction F1 to F3). This trend is confirmed after 6 months, with still reduced acido-soluble (F3: 70 ␮g g−1 ) and reducible (F4: 70 ␮g g−1 ) fractions, and without any increase of the other ones. After 12 months, a “new” zinc contamination was detected and attributed to the anode cleaning operation, performed 5 days before the final sampling (see Section 2.2). Compared to the previous sampling, higher quantity of zinc were noticed, with more zinc present in the acido-soluble fraction (F3: 141 ␮g g−1 ) and much more zinc in the reducible fraction (F4: 170 ␮g g−1 ). This quite important rise of the reducible fraction might be correlated to the release of some zinc compounds, as zinc oxides which are reducible compounds. Furthermore, results obtained on SS12 were compared to those determined on SS3, as these two samples presented quite the same total zinc content (respectively 500 ␮g g−1 and 537 ␮g g−1 ). However, zinc fractionation is quite different in these two samples: the total labile fraction (F6) is significantly lower for the sample SS12, after the anode cleaning operation (F6: 63%) compared to the SS3 samples (F6: 84%). This suggests that zinc chemical speciation is responsible for such difference, as the zinc species, resulting from the anodes cleaning (i.e., ZnO, see Section 3.1) appear clearly less available than the dissolved Zn(II) species, liberated in the solution by the anode oxidation (reaction (4)). 4. Conclusion The use of sacrificial anodes to apply cathodic protection induces a significant water contamination by Zn2+ ions. But when the Zn2+ concentration exceeds the solubility threshold (1.16 mg L−1 , at pH 8.2), this excess of free ions is consummated by precipitation (i.e., as Zn(OH)2 ) or by complexation [21]. Zinc could be also sorbed and/or

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precipitated onto the surface sediment. Zinc, coming from this permanent anode dissolution, is mainly scavenged, in the sediments, into the acido-soluble fraction. This fraction is usually related to metal bound to carbonate compounds. The contaminated sediments then appear as a secondary pollution source, as zinc can be remobilized towards water, depending on possible modifications of the physicochemical parameters, i.e., pH changes, or zinc water concentration variation. When this concentration becomes inferior to the solubility threshold, for example after a high tide exposition or during harbour dredging operations, the zinc associated to solid phases can be released into water. Such exchange between the sediment phase and the water column can induce to potential toxic effect under new environmental conditions. The anode cleaning operations (necessary to maintain an efficient protection under controlled conditions and which naturally occurs under real conditions), also induces water and superficial sediments contamination. But this phenomenon is rapidly attenuated. However, the chemical speciation of the liberated zinc might be quite different from the zinc speciation resulting from the sacrificial anodes dissolution. This is in agreement with the mineralogical fractionation results, which indicate that zinc, coming from the anodes cleaning is less available that the one resulting from the anode dissolution. Indeed, the zinc, coming from the anodes cleaning, such as zinc oxides, is mainly scavenged into the reducible fraction. Then the environmental impact of zinc coming from zinc sacrificial anodes, depends on its origin (anode cleaning/anode oxidation), but these two effects add up. This study provides information on the potential mobility of zinc in marine sediments, which depends on its chemical speciation. Under the developed experimental conditions (basin closed with low renewal), zinc contamination of water and surface sediments is higher than expected under environmental conditions, but these semi natural conditions can simulate enclosed harbour areas, which are not exposed to the tide. The concentrations observed are potentially toxic for the environment, especially for the filter feeding organisms, and could cause the whole food chain contamination. Acknowledgements This research was supported by the European Regional Development Fund (ERDF), the Conseil Régional de Basse Normandie and the Conseil Général de la Manche. The authors acknowledge the Laboratoire de Morphodynamique Continentale et Côtière of Caen-Basse Normandie University for their contributions to the sediments characterization. References [1] J. Perkins, R.A. Bornholdt, The corrosion product morphology found on sacrificial zinc anodes, Corrosion Science 17 (1977) 377–384.

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