Linda Nordmyr
email: linda.nordmyr@abo.fi

Mats Peter Österholm

Mats Åström

Estuarine behaviour of metal loads leached from coastal lowland Acid Sulphate soils

Keywords: metal load, estuarine sediments, sediment-trap material, acid sulphate soils, Vörå River, Baltic Sea 1

The estuarine behaviour of the metal load leaching from acid sulphate (AS) soils was studied in a selected river system (the Vörå River), in western Finland. Large amounts of metals were transported with the river and deposited within the estuary, causing highly elevated metal concentrations in both the sediment traps and in the underlying bottom sediments. Among the metals, there was a diverging deposition pattern where Al, Cu, La and U demonstrated a strong association with organic matter and were deposited within approximately 4 km from the river mouth. In contrast, the deposition of Co, Mn, Ni and Zn occurred when pH reached circumneutral conditions further out in the estuary. Yet other metals were not abundantly leached from the AS soils and thus not elevated in the river and estuary (Fe, Ti, Cr, V). Five separate chemical extractions indicated the geochemical speciation of the metals.

INTRODUCTION

Acid sulphate soils (AS soils) occupy worldwide an area of at least 17 million ha [START_REF] Andriesse | Acid sulfate soils: Distribution and extent[END_REF], and are formed on sulphide-rich marine sediments. Disturbance of such sediments by anthropogenic operations like draining for agricultural purposes, and to a smaller extent river dredging and excavation for road-and house building, enables oxidation of the sulphides therein. As a consequence sulphuric acid is produced and an AS soil is developed [START_REF] Dent | A world perspective on acid sulphate soils[END_REF]van Breemen, 1973). The subsequent low pH in these soils (pH <4) results in release of chemical elements from the sulphides and accelerates the mineral weathering, i.e. alteration of silicates and dissolution of associated metals (e.g. [START_REF] Joukainen | Environmental impacts and acid loads from deep sulfidic layers of two well-drained acid sulfate soils in western Finland[END_REF][START_REF] Sohlenius | Geochemistry and partitioning of trace metals in acid sulphate soils in Sweden and Finland before and after sulphide oxidation[END_REF][START_REF] Åström | Geochemistry and acidity of sulphide-bearing postglacial sediments of western Finland[END_REF][START_REF] Österholm | Spatial trends and losses of major and trace elements in agricultural acid sulphate soils distributed in the artificially drained Rintala area[END_REF].

The environmental problems related to AS soils are continuously increasing, since these soils are mainly located in coastal areas where there is an increasing pressure from population and industry, including regions such as eastern and western Australia, the Mekong Delta, western Africa, Sweden and Finland [START_REF] Andriesse | Acid sulfate soils: Distribution and extent[END_REF][START_REF] Dent | A world perspective on acid sulphate soils[END_REF]. The extensive export of acidity from these soils and the often associated biological damage, have been highlighted in many studies on several continents over the last two to three decades (e.g. [START_REF] Cook | Export of Acidity in Drainage Water from Acid Sulphate Soils[END_REF][START_REF] Dent | A world perspective on acid sulphate soils[END_REF][START_REF] Hudd | Acidification-induced Species Shifts in Coastal Fisheries off the River Kyrönjoki, Finland: A Case Study[END_REF][START_REF] Macdonald | Discharge of weathering products from acid sulfate soils after a rainfall event, Tweed River, eastern Australia[END_REF][START_REF] Minh | Contamination of surface water as affected by land use in acid sulphate soils in the Mekong delta, Vietnam[END_REF][START_REF] Ritsema | Acid sulphate soils[END_REF][START_REF] Sammut | Acidification of an Estuarine Tributary in Eastern Australia due to Drainage of Acid Sulphate Soils[END_REF][START_REF] Van Beers | Acid sulphate soils[END_REF].

In Finland, which have the largest AS soil occurrences in Europe, discharge from AS soils exports hundreds of tons of metals and acidity into coastal rivers [START_REF] Edén | Natural and land-use induced load of acidity, metals, humus and suspended matter in the Lestijoki river, W Finland[END_REF][START_REF] Roos | Hydrochemistry of rivers in an acid sulphate soil hotspot area in western Finland[END_REF][START_REF] Österholm | Quantification of current and future leaching of sulfur and metals from Boreal acid sulfate soils, western Finland[END_REF], with serious effects on many fish species that use coastal areas as a hatching ground. An acid and metal flush coinciding with the hatching period can occasionally exterminate whole fish populations [START_REF] Hudd | Bad matching between hatching and acidification: a pitfall for the burbot, Lota lota, off the river Kyrönjoki, Baltic Sea[END_REF]. In Finland several widespread mass kills of adult fish have been documented (e.g. [START_REF] Hudd | Acidification-induced Species Shifts in Coastal Fisheries off the River Kyrönjoki, Finland: A Case Study[END_REF] and a recent large one occurred in autumn 2006.

The estuarine behaviour of the metal loads leached from AS soils is poorly known. Part of the discharged metals can be lost by processes such as acid and/or humus related precipitation/sorption followed by sedimentation and also to some extent by uptake by aquatic plants in streams and rivers before entering the estuary. However, when the acid metal-rich river water enters the estuary with higher pH and salinity, the metals are likely to be removed from solution through precipitation [START_REF] Braungardt | Metal geochemistry in a mine-polluted estuarine system in Spain[END_REF][START_REF] Johnson | The regulation of trace element concentrations in river and estuarine waters contaminated with acid mine drainage: The adsorption of Cu and Zn on amorphous Fe-oxyhydroxides[END_REF][START_REF] Lopez-Gonzalez | Metal fractionation in oxic sediments of an estuary affected by acid mine drainage (south-western Spain)[END_REF][START_REF] Sholkovitz | Flocculation of dissolved organic and inorganic matter during the mixing of river water and seawater[END_REF] and consequently deposited within the estuary. Despite the massive export of metals from AS soils only a few studies have so far examined the fate of the metal load in the recipient estuaries [START_REF] Heikkilä | Human influence on the sedimentation in the delta of the river Kyrönjoki, Western Finland[END_REF][START_REF] Nordmyr | Metal pollution of estuarine sediments caused by leaching of acid sulphate soils[END_REF][START_REF] Macdonald | Impacts of runoff from sulfuric soils on sediment chemistry in an estuarine lake[END_REF].

The objective of this study, which focuses on the boreal AS soil landscape, was to: (1) identify the transport patterns of metals from the river out to the estuary, (2) determine in which forms the metals are bound in the sediment-trap material and in the bottom sediments, (3) assess the depositional mechanisms, and (4) elucidate the extent of spatial metal contamination in the estuary.

DESCRIPTION OF THE STUDY AREA

The Vörå River situated in western Finland, discharging into the Gulf of Bothnia (Fig. 1), is 35 km long and has a drainage area of 223 km 2 of which 31% is agricultural fields, 54% forest and 15% mires. No lakes exist in the drainage area and the influence of industrial and urban effluents is minor. In the river valleys the dominating Quaternary deposits are the sulphide-bearing clays and silts, which are largely used as agricultural land and wherein AS soils are developed. The precise extent of AS soils in the area is unknown, but results by [START_REF] Roos | Hydrochemistry of rivers in an acid sulphate soil hotspot area in western Finland[END_REF] indicate that the proportion of these soils is very large and most likely they exist under the main part of the agricultural land. Elsewhere, in the forested areas, glacial till (podzols) dominates. The bedrock in the area consists mainly of acidic Precambrian granites, granodiorites and gneisses [START_REF] Laitakari | Vaasa, B3. Suomen geologinen yleiskartta[END_REF] and therefore the area has an overall low buffering capacity.

[Figure 1 The inner estuary is shallow (2-10 m deep) and sheltered from wave action by surrounding land areas and islands, while further out (at Oravaisfjärden, Monåfjärden and Kalotfjärden) the water depths are commonly between 10 and 17 m (Fig. 1). The water level in the area is unaffected by tide, but affected by wind and air pressure which commonly results in ± 20 cm sea level variations and occasionally more than ± 50 cm (Fig. 2; data obtained from the Finnish Institute of Marine Research). The mean temperature in the area is 3˚C and the mean annual precipitation approximately 500 mm. During a significant part of the year (December-April/May) the river and estuary are ice covered thus limiting the runoff and transport of material to the estuary.

[Figure 2 here]

MATERIAL AND METHODS

Sampling and sample treatment

Depositing particulate matter was collected in LIMNOS sediment traps (polycarbonate plastic cylinders with d = 6.2 cm and h = 40 cm). The traps were placed at three sites along the expected river plume path in the Vörå estuary (Fig. 1; Table 1) [Table 1 here]

The material collected in the sediment traps was mostly a brown fluffy material with contributions of black and grey layers. Such layers indicate that there might exist relatively rapid variations in the sedimentation patterns within the sampling periods. After sampling, the material was immediately transported to the laboratory, where excess water was removed after the particulate matter had settled. The remaining material was dried at 40˚C and weighed. Before metal analysis, the dried material was comminuted by hand in an agate mortal. The reported metal concentrations per weight of dried samples were corrected for the weight of salt in the evaporated water, according to [START_REF] Blomqvist | Detrital bedrock elements as tracers of settling resuspended particulate matter in a coastal area of the Baltic Sea[END_REF].

At every sampling occasion (2004)(2005)(2006) the specific electric conductivity (EC), pH and temperature were measured (in situ) at the same locations as the sediment traps with a multiparametric probe (YSI model 63), in a water profile extending from the water surface down to the bottom (Fig. 1; Table 1).

Water samples (n=12) were at the same time collected from the lower part of Vörå River. The water samples were collected in polyethylene bottles and acidified with 0.2% ultrapure HNO 3 . The runoff is not monitored in the Vörå River, therefore we used the runoff from the nearby (c. 20 km away) Kyrö River (Skatila) which has similar meteorology and catchment characteristics as Vörå River. The mean specific runoff for Kyrö River (Skatila) was 8.8 l s -1 km -2 for year 2004-2006 (Fig. 2).

In September 2004 three bottom sediment profiles were sampled with a Niemistö gravity corer [START_REF] Niemistö | A gravity corer for studies of soft sediments[END_REF] at the same locations as the sediment traps. In the field, sediment cores were divided into 2-cm thick sub samples and immediately placed in airtight plastic bags. In order to minimise oxidation, the samples were kept cool until placed in a freezer. Before metal analysis, the sub samples were dried at 40˚C and comminuted by hand in an agate mortal. At site H an erosion layer (coarse material) was reached at 6 cm. Consequently, only a 6-cm long core is presented from this site. All three sediment cores had a c. 4-cm brown oxidised surface layer. At site B the 22-cm long core was black at 4-14 cm and at 14-22 cm brown-grey spots were observed in the material. At site SG the 22cm long core was black at 6-22cm. The black colour in the sediment is most likely caused by monosulphides and the grey colour indicates absence of such mineral [START_REF] Georgala | Paleoenvironmental studies of post-glacial black clays in north-eastern Sweden[END_REF]Wiklander and Hallgren, 1949). As it only takes small concentrations of monosulphides to produce the characteristically black colour [START_REF] Smith | Iron monosulfide formation and oxidation in drain-bottom sediments of an acid sulfate soil environment[END_REF], other sulphide species are likely to co-exist.

Chemical analysis

Dried and pulverized sub samples of both the sediment-trap material and bottom sediments were sent to Acme Analytical laboratories Ltd. (ISO 9001 accredited) for multi elemental analysis. In addition to conventional quality assurance/control by the lab (including calibration against standards, blanks and reruns), the analytical precision was controlled and determined from 7-20 replicates analysed in random order (see below). The concentrations of 12 metals were determined after five separate selective extractions (not sequentially), which are described below.

(1) Strong 4-acid (total) extraction: A 0.25 g portion of dried and pulverised sample was heated in a solution of HClO 4 , HNO 3 and HF to fuming and taken to dryness. The residue is dissolved in HCl and the solution is analysed with an inductively coupled plasma mass spectrometer (ICP-MS). A total of 90 samples, including blind replicates (n=20), were analysed in random order for Al (analytical precision:

1.5%), Co (2.7%), Cr (4.4%), Cu (3.2%), Fe (2.8%), La (4.8%), Mn (2.0%), Ni (3.2%), Ti (1.6%), U

(5.9%), V (4.3%) and Zn (3.5%). The analytical precision was calculated based on the random replicate pairs according to the method proposed by [START_REF] Gill | Modern Analytical Geochemistry. An Introduction to Quantitative Chemical Analysis Techniques for Earth[END_REF]. Standards and blanks were repeatedly run to control the analytical quality and the samples were analysed in random order to exclude potential artefacts due to instrumental drift. This leach dissolves the most resistant fractions and thus represents the total concentrations of the samples (Hall et al., 1996a).

( 2) Aqua regia extraction: A 0.5 g portion of the dried and pulverised sample was partially digested in 3 ml 2:2:2 HCl:HNO 3 :H 2 O (aqua regia) for one hour at 95°C, diluted to 10 ml and thereafter analysed with an ICP-MS. A total of 90 samples, including blind replicates (n=18), were analysed in random order for Al (analytical precision: 2.9%), Co (2.8%), Cr (5.3%), Cu (2.3%), Fe (2.2%), La (3.3%), Mn (4.7%), Ni (1.9%), Ti (5.6%), U (4.8%), V (4.3%) and Zn (2.5%). Aqua regia efficiently dissolves organic matter, metal-sulphides, phyllosilicates such as clay minerals, and some micas (mainly trioctahedral), while quarts, feldspars, amphiboles and dioctahedral micas (sericite or muscovite) are dissolved to a limited extent or not at all [START_REF] Chao | Use of partial dissolution techniques in geochemical exploration[END_REF][START_REF] Koljonen | Solubility in aqua regia of selected chemical elements occurring in the fine fraction of till[END_REF][START_REF] Räisänen | Relationship between mineralogy and the Physicochemical properties of till in central Finland[END_REF]. The reference data used in this study, referred to as background concentrations and representing aqua regia extractable concentrations, were received from a study by [START_REF] Åström | Geochemistry and acidity of sulphide-bearing postglacial sediments of western Finland[END_REF], where 317 samples of reduced mother sediments (deposited a long time before the industrialisation) were collected throughout mid-western Finland.

(3) Hydroxylamin (Ha) extraction: A 1 g portion of dried and pulverised sample was digested with 10 ml 0.25M hydroxylamine (0.25M NH 2 OH*HCl in 0.25M HCl) in 90°C for two hours (pH c. 1-2) and thereafter analysed by ICP-MS. A total of 90 samples, including replicates (n=19) were analysed in random order for Al (5.3%), Co (4.6%), Cu (7.2%), Fe (3.6%), La (5.3%), Mn (7.3%), Ni (5.1%), Ti (0.0%), U (3.9%), V (2.9%) and Zn (3.4%). The reason for the value of 0.0% for Ti is that for many replicate pairs the two values were identical and close to the detection limit (low resolution). For such data, the precision calculated according to [START_REF] Gill | Modern Analytical Geochemistry. An Introduction to Quantitative Chemical Analysis Techniques for Earth[END_REF] is misleadingly low. According to e.g. [START_REF] Chao | Use of partial dissolution techniques in geochemical exploration[END_REF] and [START_REF] Hall | Analytical perspective on trace element species of interest in exploration[END_REF], the Ha leach dissolves elements held in amorphous Fe-hydroxides and crystalline Mn-hydroxides. Since the samples were dried before analysis, it is likely that oxidation of monosulphides into ferric oxides and elemental S has occurred, while pyrite should have remained largely unaffected (van Breemen, 1973). Since drying of sediments promotes oxidation of Fe and Mn, the amount of Fe-and Mn-oxyhydroxides might be somewhat overestimated [START_REF] Bordas | A Critical Evaluation of Sample Pretreatment for Storage of Contaminated Sediments to be Investigated for the Potential Mobility of their Heavy Metal Load[END_REF]Zhang et al., 2001).

Based on the results of the extractions above (1-3), 17 samples from sediment traps were selected and separately extracted with sodium pyrophosphate and ammonium acetate-EDTA, according to the methods below:

(4) Sodium pyrophosphate (Na-pr) extraction: A 1.0 g portion of dried and pulverised sample was digested with 10 ml 0.1 M Na 3 P 2 O 7 (sodium pyrophosphate) and rolled for one hour, and thereafter analysed by ICP-MS. A total of 36 samples including blind replicates (n=7) were analysed in random order for Al (8.6%), Co (3.6%), Cu (33%), Fe (5.8%), La (1.8%), Mn (4.4%), Ni (8.0%), U (3.2%), V (0.7%) and Zn (2.9%). Ti was under the analytical detection limit. Sodium pyrophosphate is an alkaline complexing agent, whose effectiveness in removing the easily extractable organic matter has been attributed to its availability to chelate with Ca and trivalent metal ions (Hall et al., 1996b). Sodium pyrophosphate extracts elements adsorbed by organic matter and does not attack sulphides nor does it dissolve significant amounts of amorphous iron oxides [START_REF] Chao | Use of partial dissolution techniques in geochemical exploration[END_REF]Hall et al., 1996b).

(5) Ammonium acetate-EDTA (Am-ac) extraction: A 1.0 g portion of dried and pulverised sample was digested with 20 ml ammonium acetate-EDTA (pH=4.65), rolled for 1 hour followed by analysis with ICP-MS. A total of 36 samples including blind replicates (n=7) were analysed in random order for Al

(5.3%), Co (3.4%), Cu (1.4%), Fe (4.3%), La (3.8%), Mn (3.7%), Ni (1.1%), U (1.8%) and Zn (3.9%).

Titan and V were under the analytical detection limit. Ammonium acetate extracts exchangeable cations adsorbed by clay and elements co-precipitated with carbonates (Filgueiras et al., 2002).

It is notable that the methods are operationally defined and that the different digestions overlap each other to some extent.

Loss on ignition (LOI) was determined on both the sediment-trap material and the bottom sediments by combustion of 1.0 g of dried sample (105 °C) in a muffle oven at 500 °C for 4 hours [START_REF] Radojevi´c | Practical environmental analysis[END_REF]. In addition, in cases where we had enough material, organic carbon (OrgC) was determined by IR detection of vapour after pyrolysis at 850 °C (LECO). According to [START_REF] Ball | Loss-on-ignition as an estimate of organic matter and organic carbon in noncalcareous soils[END_REF], [START_REF] Radojevi´c | Practical environmental analysis[END_REF], [START_REF] Heiri | Loss on ignition as a method for estimating organic and carbonate content in sediments: reproducibility and comparability of results[END_REF] and [START_REF] Boyle | A comparison of two methods for estimating the organic matter content of sediments[END_REF], LOI is proportional to the concentration of OrgC and is a rapid and reasonably accurate method for estimating the organic matter concentration and can thus be used to estimate the organic matter concentration. This was also checked and verified for the bottom sediment samples of this study, in which LOI and OrgC were strongly correlated (r=1.00 with both Pearson and Spearman correlations) and LOI was higher than

OrgC by a factor of 2.7. Consequently, only LOI is discussed below.

The unfiltered and acidified water samples from Vörå River were analysed for chemical elements at Activation Laboratories Ltd. with an inductively coupled plasma mass spectrometer/optical emission spectroscope (ICP-MS/OES). In addition to conventional quality assurance/control by the lab (including calibration against standards, blanks and reruns), the analytical precision was controlled and determined from 10 replicates analysed in random order. A total of 22 samples including blind replicates (n=10) were analysed in random order for Al (0.9%), Co (3.6%), Cu (13%), Fe (2.0%), La (4.6%), Mn (1.4%), Ni (17%), Ti (5.3%), U (1.8%), V (7.4%) and Zn (7.8%). Earlier studies (e.g. [START_REF] Åström | Impact of acid sulphate soils on stream water geochemistry in western Finland[END_REF][START_REF] Åström | Abundance, Sources and Speciation of Trace Elements in Humus-Rich Streams Affected by Acid Sulphate Soils[END_REF] have concluded that metals in acid rivers like Vörå River are mostly in dissolved form. Therefore, the results in this study can be considered quite comparable to analysis of filtered samples in other studies.

Mixing experiment

An experiment with sea water mixed with 0%, 25%, 50%, 75% and 100% river water (Vörå River) was performed in order to determine the relation between EC and pH to the proportion of sea water mixed with river water. The specific EC was measured after c. 30 minutes of mixing and the results showed that EC was strongly linearly proportional to the mixing ratio. Consequently, it was possible to reliably determine the proportion of river water in the water profiles from EC measurements. The mixing ratios were also fairly consistent with pH (linear to H 3 O concentration) and therefore pH in the various part of the estuary is to a large extent a function of the mixing ratio and pH in the river.

Contaminant load

The approximate metal load transported by the Vörå River was calculated as follows:

tonne a -1 = C i * q * A
where C i is the runoff weighted average concentration (g l -1

) of element i in the Vörå River during 2004-2006, q is the mean specific runoff from Skatila for the period 1.1.2004 -31.12.2006 (8.8 l s -1 km - 2

) and A is the drainage area (223 km 2 ).

RESULTS

Hydrochemistry of Vörå River

The pH in the Vörå River was low (median 4.5) (Table 1) and uncorrelated with runoff, while the specific electric conductivity (EC) and the concentrations of Al, Co, Cu, La, Mn, Ni, U and Zn increased with the high water flows in spring and autumn (Fig. 3). Compared to the nearby Lappfjärd River with no (or minor) AS soil impact [START_REF] Roos | Hydrochemistry of rivers in an acid sulphate soil hotspot area in western Finland[END_REF], these metals were 5-72 times enriched in the Vörå River (Table 2). The concentrations of Cr, Fe, Ti and V were only marginally elevated or even depleted as compared to Lappfjärd River (Table 2) and uncorrelated with flow (Fig. 3). The mean annual load from the Vörå River drainage area was in decreasing order 590 tonne (t) Al > 110 t Mn > 100 t Fe > 13 t Zn > 6.4 t Ni > 3.8 t La > 3.0 t Co > 1.9 Cr > 0.85 t Cu > 0.79 Ti > 0.75 t U > 0.60 t V.

[Figure 3 and table 2 here]

Hydrochemistry of the estuary

Based on EC the layer of lighter fresh river water was as expected thickest in the vicinity of the river surface water pH at site H and B was occasionally low but was for the most part above 7 (Fig. 4) and were consistent with the pH of the river (Fig. 3). However, the runoff and pH at the different sites were also very consistent i.e. the lowest pH at these sites is found at high runoff (Fig. 4; Fig. 2). At site SG, the pH was considerably lowered in autumn 2006 at the highest runoff, but otherwise variations in pH and runoff did not affect pH much at this site (Fig. 4).

[Figure 4 here] The amount of settling material was as expected highest closest to the river mouth at site H and lowest in the outermost estuary at site SG (Fig. 5). At site B the lower trap collected somewhat more material than the upper trap (Fig. 5). The variations in deposition were relatively constant at site B and SG, while significant fluctuations occurred at site H (Fig. 5). There were only weak indications (not statistically significant) that runoff increased the overall amounts of deposited material and metals (Fig. 5). Since the deposition under ice was only determined once in March-May 2006 at site H, the amount of material deposited during the winter period is poorly known. However, the data indicates minimal deposition in the drainage area during winter (Fig. 5). The sedimentation rate based on trap measurements (Fig. 5) is in reasonable good agreement with corresponding sedimentation rates obtained by 137 Cs dating on sediment cores in a previous study by [START_REF] Nordmyr | Metal pollution of estuarine sediments caused by leaching of acid sulphate soils[END_REF].

[Figure 5 here]

Chemistry

Based on the temporal and spatial variations in the aqua regia extractable elemental concentrations in the trap material, the metals can be organized in three groups: (1) those that occurred below background concentrations (Fe, Cr, Ti and V) (Fig. 6a), (2) those that were enriched at site H and B

(1.7 km and 4 km from the river mouth, respectively) and occurred close to background concentrations (Al, Cu, La and U) at site SG (c. 9 km from the river mouth) (Fig. 6b) and ( 3) those that were particularly strongly enriched in both the upper and lower trap at site B (Co, Mn, Ni and Zn) (Fig. 6c).

These results are thus in line with the hydrochemical data for the inflowing river, with low relative concentration of Cr, Fe, Ti and V and elevated relative concentrations of Al, Co, Cu, La, Mn, Ni and Zn (Fig. 3; Table 2). The organic matter (LOI) concentration in the trap material had a pattern similar to the metals of the second group (Fig. 6b). In terms of aqua regia concentrations of trap material, the metals of group 2 tended to increase with decreasing pH in the estuary, while concentration of metals in group 3 tended to increase with increasing pH (Fig. 4; Fig. 6b-c). There was no clear connection between the overall metal discharge from the river and corresponding metal deposition in the sediment traps.

[Figure 6a-c here]

The concentrations of metals extracted by the weaker reagents showed only small variations over time (Fig. 7a-c). Since extractions by Am-ac and Na-pr were only done on selected samples we preferred to illustrate these results with median, minimum and maximum values only (Table 3).

The Ha leach efficiently extracted (medians in decreasing order) Mn (c. 85% of the aqua regia concentrations) > Zn (75%) > Co (51%) > La (48%) > Ni (48%) > Fe (43%) > Al (40%) > U (32%). This

indicates that these metals are abundantly associated with Al-, Fe-and Mn-oxyhydroxides and/or possibly to some degree to monosulphides and carbonates. Copper, Ti and V were on the other hand weakly extracted with Ha (Table 3).

The Na-pr leach efficiently extracted U (c. 53% of the aqua regia concentrations) and moderately to weakly (in decreasing order) Zn (23%) > Cu (19%) > La (18%) > Al (14%) > Mn (13%) > Ni (8%) > Fe (8%) > V (7%) > Co (6%).

The Am-ac leach efficiently extracted U (c. 46% of the aqua regia concentrations) and Mn (30%), and moderately to weakly (in decreasing order) Zn (21%) > La (14%) > Al, Co, Ni (9%) > Cu (8%) > Fe (2%).

[Table 3 here]

[Figure 7a-c here]

Bottom sediments

The concentrations and spatial patterns of the aqua regia extractable metal concentrations and LOI in the bottom sediments were very similar to those in the corresponding sediment-trap materials (Fig. 6ac). In addition, the different chemical extractions delivered approximately similar results for the bottom sediments as for the sediment-trap material, i.e. the metals are associated to the same fractions in both the sediment-trap material and in the bottom sediments (Fig. 7a-c).

At site SG (furthest out in the estuary) the aqua regia extractable concentrations of Co, Mn, Ni and Zn were higher in the trap material than in the underlying bottom sediments (Fig. 6a-c). In addition, Mn and associated metals (Co, Ni and Zn) were to a large part associated with the Ha fraction in the uppermost oxidized layer in the bottom sediment at this site (Fig. 7c). At site B, an opposite pattern was seen, i.e. only small amounts of Mn and the associated metals (Co, Ni and Zn) were extracted with Ha in the surface sediment, while the Ha fraction increased with depth (Fig. 7c).

DISCUSSION

Cr, Fe, Ti & V

Chromium, Fe, Ti and V were not enriched in any of the media examined in this study and are consequently not particularly mobile in AS soil landscapes (Table 2; Fig. 6a). The background concentrations of Cr, Ti and V in the river water and estuarine sediments are explained by their participation in chemically resistant fractions in the AS soils [START_REF] Sohlenius | Geochemistry and partitioning of trace metals in acid sulphate soils in Sweden and Finland before and after sulphide oxidation[END_REF][START_REF] Åström | Partitioning of transition metals in oxidised and reduced zones of sulphide-bearing fine-grained sediments[END_REF]. The low relative mobility and transport of Fe are typical for Finnish (boreal) AS soils [START_REF] Nordmyr | Estimation of leakage of chemical elements from boreal acid sulphate soils[END_REF][START_REF] Sohlenius | Geochemistry and partitioning of trace metals in acid sulphate soils in Sweden and Finland before and after sulphide oxidation[END_REF][START_REF] Åström | Partitioning of transition metals in oxidised and reduced zones of sulphide-bearing fine-grained sediments[END_REF][START_REF] Österholm | Spatial trends and losses of major and trace elements in agricultural acid sulphate soils distributed in the artificially drained Rintala area[END_REF] and a possible explanation is precipitation of Fe-oxyhydroxides in soil cracks after oxidation of sulphides, thus inhibiting release of Fe from these soils [START_REF] Ritsema | The transformation of potential into actual acid sulphate soils studied in column experiments[END_REF][START_REF] Sohlenius | Geochemistry and partitioning of trace metals in acid sulphate soils in Sweden and Finland before and after sulphide oxidation[END_REF].

In a study by [START_REF] Åström | Abundance, Sources and Speciation of Trace Elements in Humus-Rich Streams Affected by Acid Sulphate Soils[END_REF] in a nearby similar river (Petalax River), Fe was mostly transported in the particulate phase as a result of co-flocculation with humic substances (>50%) and only to a minor extent as Fe-oxyhydroxides coated with negatively charged humic substances (the anionic fraction) and as Fe 2+ and/or uncoated positively charged Fe-oxyhydroxides (cationic fraction).

In the trap material and in the bottom sediments Fe was to a large extent extracted with Ha (c. 40 % of the aqua regia concentration), which shows that Fe abundantly occurs as hydroxides and/or monosulphides (Table 3). Because none of the metals enriched in the studied materials had a distribution pattern similar to Fe, it is assumed that Fe-oxyhydroxides are of minor importance in controlling the geochemical patterns and dynamics in the studied environment.

Al, Cu, La & U

The high concentrations and loads of Al, Cu, La and U (the Al-group metals) in the river water (Table 2; Fig. 3) are a result of leakage from the AS soils in the area [START_REF] Åström | Impact of acid sulphate soils on stream water geochemistry in western Finland[END_REF], since the low pH in these soils cause phyllosilicates to release large amounts of metals like Al [START_REF] Ritsema | The transformation of potential into actual acid sulphate soils studied in column experiments[END_REF]van Breemen, 1973) and La into the drainage water. Copper on the other hand is most likely released from the sulphides (Huerta-Diaz and [START_REF] Huerta-Diaz | Pyritization of trace metals in anoxic marine sediments[END_REF] and U from oxidation of U (IV) minerals in the AS soils (Åström et al., accepted). In the nearby AS soil affected Petalax River, sampled on one occasion by [START_REF] Åström | Abundance, Sources and Speciation of Trace Elements in Humus-Rich Streams Affected by Acid Sulphate Soils[END_REF], Al was mostly transported in cationic form (Al 3+ , inorganic phase) but also to some extent in the humus-associated phase (anionic form) and on particles. On the other hand, 50% of Cu was transported as particles and in humus-associated phases (anionic form) and 50% in the inorganic phase (cationic form) [START_REF] Åström | Abundance, Sources and Speciation of Trace Elements in Humus-Rich Streams Affected by Acid Sulphate Soils[END_REF]. La and U were not included in that study.

The similar spatial and temporal concentration patterns of LOI and the Al-group metals (aqua regia extractable) in both the sediment-trap material and the underlying bottom sediments (Fig. 6b) indicate that these metals are strongly associated with organic matter. This is supported by the association of Al and Cu with humic substances in the river (see above), and in line with results of other studies [START_REF] Boyle | The mechanism of iron removal in estuaries[END_REF][START_REF] Elbaz-Poulichet | Trace metal and nutrient distribution in an extremely low pH (2.5) river-estuarine system, the Ria of Huelva (South-West Spain)[END_REF][START_REF] Forsberg | Trace Metal Speciation in Brackish Water Using Diffusive Gradients in Thin Films and Ultrafiltration: Comparison of Techniques[END_REF][START_REF] Hoyle | The behaviour of the rare earth elements during mixing of river and sea waters[END_REF][START_REF] Sholkovitz | Flocculation of dissolved organic and inorganic matter during the mixing of river water and seawater[END_REF]Widerlund, 1996). In addition, the Al-group metals were moderately to extensively extracted with Na-pr (removes easily extractable organic matter), i.e. 14-53% of the aqua regia extractable concentration (Table 3). The relatively low extractability of Cu with both Na-pr and Ha is noteworthy (Table 3), and suggests that Cu is preferentially associated with an organic-matter phase not dissolved by the Na-pr extraction (Hall et al., 1996b).

Aluminium and La were, however, more abundantly leached with Ha than with Na-pr (Table 3). This is interpreted as association preferentially with a hydroxide phase, which is in line with our field observations of lighter coloured water (even whitish) in the inner estuary compared to the river mouth and with flocs clearly visible approximately within 200 m from the river mouth. In Australia the same phenomenon with white flocs has occasionally been observed and has been interpreted as Alhydroxides or Al-hydroxysulfate [START_REF] Sammut | Acidification of an Estuarine Tributary in Eastern Australia due to Drainage of Acid Sulphate Soils[END_REF]. The importance of precipitation/adsorption processes in the estuarine zone, where a sharp increase in pH and salinity occurs, is well documented (e.g. [START_REF] Boyle | The mechanism of iron removal in estuaries[END_REF][START_REF] Crerar | Hydrogeochemistry of the New Jersey coastal plain[END_REF][START_REF] Johnson | The regulation of trace element concentrations in river and estuarine waters contaminated with acid mine drainage: The adsorption of Cu and Zn on amorphous Fe-oxyhydroxides[END_REF]. Since the Al-group metals have Copper is to a lesser extent transported to site B compared to Al, La and U (Fig. 6b). This indicates that Cu is more readily precipitated close to the river mouth than the other metals. Lanthanum on the other hand occurred in relatively high concentrations at site B, as compared to those at site H (Fig. 6b). This may be pH dependent patterns, as it can be expected that pH always increase, more or less, when going from the river mouth into the estuary.

In autumn 2006, after a extremely dry summer followed by abundant rain, the Vörå River and its estuary were strongly affected by acid and metal rich waters discharging from the AS soils (river water sampling was done in November; Fig. 3), resulting in serious and significant fish kills and turquoise coloured water as far out as 7-8 km from the river mouth. At this time at site H, thick layers of Al-flocks were observed at the bottom and significant amounts of dead fish. However, at site SG no Alhydroxide precipitates were observed, which is in line with the chemical data (Fig. 7b). The fish probably died as a result of low pH and Al-precipitates clogging the gills [START_REF] Bjerknes | Aluminium in acidic river water causes mortality of farmed Atlantic Salmon (Salmo salar L.) in Norwegian fjords[END_REF][START_REF] Peuranen | Effects of acidity and aluminium on fish gills in laboratory experiments and in the field[END_REF][START_REF] Björkqvist | Liming as a method to neutralize highly acidic drainage waters from sulphate basins in western Finalnd[END_REF]. While there is no question that the huge amounts of metals in the river in autumn 2006 (Fig. 3) were caused by the flushing of farmland AS soils in the catchment, sulphidic dredge materials excavated in autumn/winter 2005 and 2006 and placed along the banks of the lower reaches of the river, may have contributed to the metal load of the river.

Co, Mn, Ni & Zn

The high concentrations and loads of Co, Mn, Ni and Zn (the Mn-group metals) in the river water (Table 2; Fig. 3) are most likely partly released from silicates in AS soils due to the low pH in the soil [START_REF] Sohlenius | Geochemistry and partitioning of trace metals in acid sulphate soils in Sweden and Finland before and after sulphide oxidation[END_REF] and from sulphides due to oxidation (e.g. Huerta-Diaz and [START_REF] Huerta-Diaz | Pyritization of trace metals in anoxic marine sediments[END_REF][START_REF] Huerta-Diaz | Geochemistry of trace metals associated with reduced sulfur in freshwater sediments[END_REF][START_REF] Morse | Chemical influences on trace metal-sulfide interactions in anoxic sediments[END_REF][START_REF] Sohlenius | Geochemistry and partitioning of trace metals in acid sulphate soils in Sweden and Finland before and after sulphide oxidation[END_REF]. These metals are presumably carried as cations in the river water, as in the nearby AS soil affected Petalax River [START_REF] Åström | Abundance, Sources and Speciation of Trace Elements in Humus-Rich Streams Affected by Acid Sulphate Soils[END_REF].

The Mn-group metals behave in a different manner than the Al-group metals in the estuary. This is seen by the much higher concentrations (aqua regia extraction) in the sediment-trap material and underlying bottom sediments at site B compared to site H (Fig. 6c). In addition, these metals are only marginally extracted by Na-pr (Table 3), which suggests that the precipitation and deposition mechanisms are different as compared to that of the Al-group metals. This is in line with the deposition patterns of Mn in an adjacent AS soil affected estuary [START_REF] Heikkilä | Human influence on the sedimentation in the delta of the river Kyrönjoki, Western Finland[END_REF], and with other studies suggesting low association of Mn and Zn with organic complexes [START_REF] Burton | Factors Controlling the Geochemical Partitioning of Trace Metals in Estuarine Sediments[END_REF][START_REF] Forsberg | Trace Metal Speciation in Brackish Water Using Diffusive Gradients in Thin Films and Ultrafiltration: Comparison of Techniques[END_REF][START_REF] Tribovillard | Trace metals as paleoredox and paleoproductivity proxies: An update[END_REF].

Manganese was nearly quantitatively extracted with the Ha leach (Fig. 7c) and moderately extracted with the Am-ac leach (Table 3). This strongly indicates that the high Mn concentrations are predominantly the result of precipitation and deposition of Mn-oxyhydroxides with additional contribution by Mn-carbonates (the acid Am-Ac leach). However, it cannot be excluded that a small amount of Mn-sulphides and/or Mn adsorbed to FeS was converted into hydroxides during sample handling. Since Co, Ni and Zn had similar spatial patterns and response to chemical reagents as Mn, it is reasonable to suggest that these metals are adsorbed and controlled by Mn-oxyhydroxides and possibly Mn-carbonates [START_REF] Drever | The geochemistry of natural waters[END_REF][START_REF] Tribovillard | Trace metals as paleoredox and paleoproductivity proxies: An update[END_REF].

The most likely explanation for the relatively low concentrations of the Mn-group metals at site H is that above circumneutral (alkaline) pH conditions are needed for significant amounts of Mn to precipitate [START_REF] Braungardt | Metal geochemistry in a mine-polluted estuarine system in Spain[END_REF][START_REF] Ponter | Geochemistry of manganese in the Kalix River, northern Sweden[END_REF]. Thus, the plume of strongly to slightly acidic surface water acts as an efficient carrier of the Mn-group metals far out in the estuary. The variable concentrations of Mn (and partly for Co, Ni and Zn) in the sediment-trap material and in the bottom sediments, particularly at site B (Fig. 6c), can consequently be ascribed to pH variations caused by the relative contribution by, and quality of, the inflowing river water. This spatial and temporal variability is supported by the geochemistry of the underlying sediments and other deep sediment cores from the same area [START_REF] Nordmyr | Metal pollution of estuarine sediments caused by leaching of acid sulphate soils[END_REF].

At site SG, the metal concentrations (aqua regia extractable) were higher in the trap material than in the underlying bottoms sediments (Fig. 6c). This indicates relatively little deposition at this site and/or resuspension due to strong flows at this relatively open and unsheltered site. In the surface layer of the bottom sediment at site SG, the Mn-group metals were, however, above background concentrations and quantitatively extracted with Ha, while down core they approached background concentrations and were only marginally extracted by this reagent (Fig. 6c; Fig. 7c). This suggests reduction of Mn-oxyhydroxides at depth and associated release of Mn 2+ into the pore water with subsequent migration up to the oxidised brownish surface layer, where re-precipitation as oxyhydroxides occurs. This mechanism is supported by other studies [START_REF] Morford | Diagenesis of oxyanions (V, U, Re, and Mo) in pore waters and sediments from a continental margin[END_REF]van den Berg et al., 2000;Widerlund and Ingri, 1996).

In contrast, at site B where the Mn-group metals were particularly abundant in both the sediment-trap material and in underlying bottom sediments (Fig. 6c), the Ha-extractable fraction in the uppermost sediment layer was lower than down the core and also lower than in the upper layers at site SG (Fig. 7c). This might be explained by a decline in bottom-water (and thus sediment) redox potential during the last few years (or during the winter before the sediment sampling), causing the release of Mn 2+ and associated metals in the surface sediments and thus leaching into the overlying water [START_REF] Morford | Diagenesis of oxyanions (V, U, Re, and Mo) in pore waters and sediments from a continental margin[END_REF]van den Berg et al., 2000). Re-precipitation of Mn-oxyhydroxides is likely to occur at oxidising conditions in the water column [START_REF] Ingri | Chemistry of suspended particles in the southern Baltic Sea[END_REF]) and thus there is potential for re-deposition of Mn-oxyhydroxides in the sediment traps.

CONCLUSION

This study shows that AS soils, formed as a result of farmland drainage, strongly affect both the chemistry of the downstream waters and the downstream estuarine sediments.

In the estuary, a strong fractionation in depositional patterns of a suite of metals abundantly leached from AS soils was detected: (1) precipitation and deposition of Al, Cu, La and U occurred in the inner estuary either together with organic matter and/or Al-hydroxides, and (2) deposition of Co, Mn, Ni and

Zn occurred in the central/outer estuary mainly in the form of Mn-oxyhydroxides with co-precipitated Co, Ni and Zn. Yet other metals (Fe, V, Cr and Ti) were not, in relative terms, abundantly leached from AS soils and thus not enriched in the river and its estuary. These patterns are largely the result of pH and salinity gradients, the behaviour of organic matter, redox variations and different precipitation and settling mechanisms

The on-going sedimentation and isostatic land uplift continuously decrease the water depth in the estuary of the Vörå River. This will ultimately lead to more pronounced pH surges, and also to extending dredging activities in order to maintain water depths required for the boat traffic. published in [START_REF] Åström | Geochemistry and acidity of sulphide-bearing postglacial sediments of western Finland[END_REF]. Since LOI and U were not determined by [START_REF] Åström | Geochemistry and acidity of sulphide-bearing postglacial sediments of western Finland[END_REF] the background values for these are from a similar study by [START_REF] Nordmyr | Estimation of leakage of chemical elements from boreal acid sulphate soils[END_REF]. is the median concentration of 317 samples of similar sediments deposited before industrialisation [START_REF] Åström | Geochemistry and acidity of sulphide-bearing postglacial sediments of western Finland[END_REF].

FIGURE CAPTIONS

  here] Vörå River is one of the most heavily AS soil-impacted rivers in Finland[START_REF] Roos | Hydrochemistry of rivers in an acid sulphate soil hotspot area in western Finland[END_REF] and Europe, and discharges into a shallow and biologically sensitive brackish water estuary. The existence of fish and other organisms is therefore very limited in these waters. Vörå River was seriously flooded in summer of 2004, causing large damages to the land areas close to the river. As a consequence the river was cleared and widened in late autumn/early winter of 2005/6 and 2006/7, without consideration of chemical effects on the environment.

  : Hellnäs (H) (water depth 10 m/distance from river mouth 1.7 km), Bentasholm (B) (17 m/4 km) and Stora Granskär (SG) (10 m/9 km). The traps were moored approximately 2 m from the bottom, and at site B an additional upper trap was placed 5 m above the bottom-trap (7 m from the bottom). The collection of particulate matter started every year shortly after the ice broke up. The traps were emptied with one-two month intervals and removed in autumns before the build up of ice. The study period thereby comprised spring, summer and autumn, and are presented in Figure 2. During the flooding year 2004, no sampling of sediment-trap material was done during the maximum flooding. However, during the dredging activity of the Vörå River in early spring (March) and in autumn (September-November) 2006, sampling of sediment-trap material was carried out.

  mouth and attenuated towards the sea. At site H (c. 2 km from the river mouth) the uppermost 2-3 m of the water column was composed of 30-80% river water and from 3 m down to the bottom (10 m) the proportion of river water was 5-30%. At site B (c. 4 km from the river mouth) the proportion of river water was somewhat lower. The uppermost 2-3 m was for the most part composed of 20-50% river water (80% due to high runoff and/or low sea level in spring 2005 (Fig.2)) and from 3 m to the bottom, the proportion of river water was consistent (5-10%). Even at site SG, which is 9 km from the river mouth, river water was detected. However the proportion of river water was well below 10% in the uppermost 3-5 m and insignificant below 5 m.Surface waters have the lowest pH, which increases downwards due to the lower proportion of river water at depth. Consequently, pH at the surface represents the minimum pH of the water profiles. The

  background concentrations at site SG, these metals are normally not transported to such far distances in the estuary. Moreover, as the highest concentrations of the Al-group metals in the sediment trap material were found during low pH conditions at site H and B (Fig. 4; Fig 6b) it is also possible that at high pH conditions a significant part of the Al is deposited (depleted from the water column) before reaching site H and B. The similar distribution pattern of Al and LOI is thus caused by either: (1) direct binding of Al by organic matter ultimately depositing in the estuary, (2) a common spatial settling pattern for Aloxyhydroxides and organic matter or (3) formation of mixed particles composed of Al-oxyhydroxides and organic compounds. Accordingly, the metals Cu, La and U, which were closely associated with LOI and Al, might either be bound to the organic matter, to the fresh Al-hydroxide precipitates or to the mixed particles.

Fig. 1 .Fig. 2 .Fig. 3 .

 123 Fig. 1. The sampling locations in the estuary of Vörå River. The shaded areas in Finland represent the land area covered by water since the retreat of the Pleistocene ice sheet. In depressions and valleys within this area, sulphur-rich brackish-water sediments are abundant on which AS soils are developed. Fig. 2. The specific runoff in Kyrö River during year 2004-2006 and for the specific sampling periods and the sea level in the nearby city of Vasa during 2004-2006. Sea level data was obtained from the Finnish Institute of Marine Research.Fig. 3. Hydrochemistry and specific runoff for the Vörå River during 2004-2006.

Fig. 4 .

 4 Fig. 4. The pH in the surface water at respective sites during the beginning and end of the sampling periods. The pH in the surface water represent the minimum pH in the water profiles.

Fig. 5 .

 5 Fig. 5. The amount of material deposited in the sediment traps and the mean specific runoff (Vörå River) during the sampling periods. Note that winter/spring sampling (17.03-24.04.2006) only occurred at site H.Fig. 6a. Aqua regia extractable concentrations of Fe, Cr, Ti and V in the sediment-trap material over time (to the left) and in the underlying bottom sediments (to the right). The background concentrationis the median concentration of 317 samples of similar sediments deposited before industrialisation[START_REF] Åström | Geochemistry and acidity of sulphide-bearing postglacial sediments of western Finland[END_REF].

Fig. 6b .

 6b Fig. 6b. Aqua regia extractable concentrations of LOI, Al, Cu, La and U in the sediment-trap material over time (to the left) and in the underlying bottom sediments (to the right). The background concentration for Al, Cu and La is from 317 sampled reduced mother sediments from western Finland

Fig. 6c .

 6c Fig. 6c. Aqua regia extractable concentrations of Co, Mn, Ni and Zn in the sediment-trap material over time (to the left) and in the underlying bottom sediments (to the right). The background concentration

Fig.

  Fig. 7a-c. The vertical distribution of different fractions of metals in the sediment cores (to the left) and the temporal and spatial distribution of the same fractions in the sediment-trap material (to the right) at site H, B and SG. Ha = Hydroxylamine extractable concentration, AR = the aqua regia extractable concentrations minus the Ha extractable concentrations, and residual = total concentrations minus the aqua regia extractable concentrations.

  There is thus a risk of both chemical and physical mobilisation of the high metal concentrations found in the bottom sediments throughout the estuary. It is therefore strongly recommended that detailed risk assessments are carried out in the Vörå River estuary, as well as in other similar estuaries in Finland and elsewhere, where an abundance of acid sulphate soils in the catchments are delivering large amounts of acidity and metals to downstream waters.Institute for LandReclamation and Improvement Publication, pp. 66-130. van den Berg, G., A., Loch, G., van der Heijdt, L., M. , Zwolsman, J., J.G., 2000. Geochemical behaviour of trace metals in freshwater sediments. In: Markert, B., Friese, K. (Eds.), Trace Metals in the Environment.Elsevier,

	Widerlund, A., 1996. Early diagenetic remobilization of copper in near-shore marine sediments: a
	quantitative pore-water model. Marine Chemistry, 54 (1-2), 41-53.
	Widerlund, A., Ingri, J., 1996. Redox cycling of iron and manganese in sediments of the Kalix River
	estuary, Northern Sweden. Aquatic Geochemistry (Historical Archive), 2 (2), 185-201.
	Wiklander, L., Hallgren, G., 1949. Studies in Gyttja Soils: I. Distribution of Different Sulfur and
	Phosphorus Forms and of Iron, Manganese, and Calcium Carbonate in a Profile from Kungsängen.
	The Annals of the Royal Agricultural College of Sweden, 16, 811-827.
	Zhang, S., Wang, S., Shan, X.-Q., 2001. Effect of sample pretreatment upon the metal speciation in
	sediments by a sequential extraction procedure. Chemical Speciation and Bioavailability, 13 (3), 69-
	74.

Table 1 .

 1 TABLES General characteristics of the sampling sites.

	Vörå River	Site H	Site B	Site SG

Table 2 .

 2 Hydrochemistry of Vörå River (VR) and Lappfjärd River (LR). The latter is used as a reference for a river not affected by AS soils.

					LR	*	VR/LR
			Median	Mean	Mean	Mean
	Al mg l -1	5.9	6.8	0.31	22
	Cu g l -1	11	11	2.5	4.5
	La g l -1	39	48	_		_
	U	g l -1	0.75	0.69	0.10	7.0
	Co g l -1	34	38	0.52	72
	Mn g l -1	1400	1500	44	34
	Ni	g l -1	77	85	2.0	42
	Zn g l -1	170	180	15	12
	Cr	g l -1	1.9	2.1	1.1	1.8
	Fe g l -1	1100	2100	770	2.8
	Ti	g l -1	3.7	13	_		_
	V	g l -1	0.60	1.3	0.84	1.5
	_ = not determined

* Data published in

[START_REF] Roos | Hydrochemistry of rivers in an acid sulphate soil hotspot area in western Finland[END_REF]

. VR

Table 3 .

 3 The elemental concentrations (median, min and max) in the sediment-trap material measured

	752
	with five different separately performed chemical extractions.

n = number of samples analysed _ = not analysed BD = bellow detection limit Total extraction (HClO 4 -HNO 3 -HCl-HF) Aqua Regia extraction (2:2:2, HCl:HNO 3 :H 2 O) Ha = hydroxylamin extraction Na-pr = Sodium pyrophosphate extraction Am-ac = ammonium acetate-EDTA extraction

van Breemen, N., 1973. Soil forming processes in acid sulphate soils. In: Drost, H. (Ed.), Acid Sulphate Soil. Proceedings of the international symposium on acid sulphate soils. International

Fig.1

Fig.2

Fig.3

Fig.4

Fig.5

Fig.6(a)

Fig.6(b)

Fig.6(c)

Fig.7(a)

Fig.7(b)

Fig.7(c)

ACKNOWLEDGMENT

This study were financially supported by Magnus Ehrnrooths Stiftelse, Åbo Akademi Forskningsinstitut, Renlunds Stiftelse, Sohlbergska delegationen and Svenska Litteratursällskapet (Ingrid, Margit och Henrik Höijers donationsfond II). We would like to thank the personnel (especially Hans-Göran Lax, Hannu Jerohin and Juha Ojaharju) at the West Finland Regional Environmental Centre, for assistance with sampling. We also would like to thank the Finnish Institute of Marine

Research for the sea level data obtained.


Stefano Vaselli

Fabio Bulleri
email: fbulleri@biologia.unipi.it

Lisandro Benedetti-Cecchi

HARD COASTAL-DEFENCE STRUCTURES AS HABITATS FOR NATIVE AND EXOTIC ROCKY-BOTTOM SPECIES

Keywords: Coastal zone, Artificial habitats, Coastal-defence, Wave-exposure, Biological invasions, Caulerpa racemosa, Codium fragile ssp. tomentosoides

The use of hard coastal-defence structures, like breakwaters and seawalls, is rapidly increasing to prevent coastal erosion. We compared low-shore assemblages between wave-protected and wave-exposed habitats on breakwaters along a sandy shore of Tuscany (North-Western Mediterranean). Assemblages were generally characterized by a low diversity of taxa, with space monopolized by Mytilus galloprovincialis and Corallina elongata on the seaward side of breakwaters and by filamentous algae on the landward side. Assemblages in wave-protected habitats were characterized by greater temporal stability than those in exposed habitats and supported non-indigenous macroalgae such as Caulerpa racemosa and Codium fragile ssp.

tomentosoides. Hence, the introduction of hard coastal defence-structures in otherwise soft-bottom dominated areas, attracting native and exotic rocky bottom species, should be of great concern for the conservation of marine biodiversity at local and regional scales and for the management of biological invasions.

Introduction

More than 60% of the human population is living in coastal zones and projections for the next decades foresee a further increase, with this fraction exceeding 75%, as a consequence of the ongoing demographic expansion and tendency to migrate towards coastal areas [START_REF] Eea | Coastal and marine zones[END_REF][START_REF] Airoldi | Loss, status, and trends for coastal marine habitats of Europe[END_REF]. The need to protect the coast from erosion and flooding have increased the use of man-made structures like breakwaters, groynes and seawalls [START_REF] Bulleri | Is it time for urban ecology to include the marine realm?[END_REF][START_REF] Airoldi | Loss, status, and trends for coastal marine habitats of Europe[END_REF]. A further increment in the presence of hard coastal-defence structures is predicted to happen in response to sea level rise and to increases in the intensity and frequency of storms [START_REF] Bray | Prediction of soft-cliff retreat with accelerating sea level rise[END_REF][START_REF] Valiela | Global Coastal Change[END_REF]. Nonetheless, in contrast to terrestrial environments, artificial marine habitats have received little attention by ecologists and our understanding of their ecological value and functioning remains limited [START_REF] Bulleri | Is it time for urban ecology to include the marine realm?[END_REF]. Only in recent years have the importance and advantages of including ecological criteria into the design and management of man-made structure been recognized [START_REF] Glasby | Differences between subtidal epibiota on pier pilings and rocky reefs at marinas in Sydney, Australia[END_REF][START_REF] Glasby | Urban structures as marine habitats[END_REF][START_REF] Davis | Artificial armored shorelines: sites for opencoast species in a southern California Bay[END_REF][START_REF] Bacchiocchi | Distribution and dynamics of epibiota on hard structures for coastal protection[END_REF][START_REF] Chapman | Paucity of mobile species on constructed seawalls: effects of urbanization on biodiversity[END_REF]Airoldi et al., 2005a;[START_REF] Moreira | Maintenance of chitons on seawalls using crevices on sandstone blocks as habitat in Sydney Harbour, Australia[END_REF].

Much of the research in this field has been done on artificial structures deployed on hardbottoms. A number of studies has compared benthic assemblages on different types of artificial habitats [START_REF] Connell | Do urban structures influence local abundances and diversity of subtidal epibiota? A case study from Sydney Harbour, Australia[END_REF][START_REF] Connell | Urban structures as marine habitats: an experimental comparison of the composition and abundance of subtidal epibiota among pilings, pontoons and rocky reefs[END_REF] or have assessed the extent to which artificial structures can be considered as surrogates of natural rocky habitats, focusing on assemblages of algae, invertebrates or fish (Connel and [START_REF] Connell | Do urban structures influence local abundances and diversity of subtidal epibiota? A case study from Sydney Harbour, Australia[END_REF][START_REF] Glasby | Differences between subtidal epibiota on pier pilings and rocky reefs at marinas in Sydney, Australia[END_REF][START_REF] Davis | Artificial armored shorelines: sites for opencoast species in a southern California Bay[END_REF][START_REF] Chapman | Paucity of mobile species on constructed seawalls: effects of urbanization on biodiversity[END_REF][START_REF] Chapman | Intertidal seawalls-new features of landscape in intertidal environments[END_REF][START_REF] Guidetti | Fish assemblages associated with coastal defence structures in south-western Italy (Mediterranean Sea)[END_REF][START_REF] Chapman | Intertidal seawalls as habitats for molluscs[END_REF][START_REF] Clynick | Assemblages of fish associated with coastal marinas in north-western Italy[END_REF]. To our knowledge, very few studies have instead assessed the ecological implications of introducing hard coastal-defence structures into areas where soft-bottoms are predominant (but see Davis et al., 2002, Bacchiocchi and[START_REF] Airoldi | The effects of sedimentation on rocky coast assemblages[END_REF][START_REF] Moschella | Low-crested coastal defence structures as artificial habitats for marine life: using ecological criteria in design[END_REF][START_REF] Pinn | The assemblages on groynes in relation to substratum age, aspect and microhabitat[END_REF][START_REF] Gacia | Low crested coastal defence structures on the Catalan coast of the Mediterranean Sea: how they compare with natural rocky shores[END_REF].

Impacts caused by their deployment have been mainly evaluated in terms of the changes caused to assemblages living in the surrounding sedimentary habitats. Changes in water flow, illumination and rates of sedimentation due to the introduction of built structures can be, in fact, detrimental for plants (e.g. seagrasses) and animals living in or on soft bottom substrata [START_REF] Davis | Man-made structures on marine sediments: effects on adjacent benthic communities[END_REF][START_REF] Barros | The influence of rocky reefs on the structure of benthic macrofauna in nearby soft-sediments[END_REF][START_REF] Martin | Ecological impact of coastal defence structures on sediments and mobile infauna: evaluating and forecasting consequences of unavoidable modifications of native habitats[END_REF]. Less attention has, in contrast, been given to the implications these structures may have in attracting a suite of rocky bottom organisms [START_REF] Bulleri | The introduction of artificial structures on marine soft-and hard-bottoms: ecological implications of epibiota[END_REF]; although this function is widely acknowledged for artificial reefs that are actually built with the primary objective of functioning as aggregating devices or as tools for the rehabilitation of endangered or overexploited species [START_REF] Collins | Coastal structures, waste materials and fishery enhancement[END_REF][START_REF] Carr | Artificial reefs: the importance of comparisons with natural reefs[END_REF]. Provision of hard substrata by manmade structures can facilitate the expansion of a number of hard-bottoms species, including those that are non-indigenous, in areas that otherwise lack suitable habitats (e.g. natural hard-bottoms) [START_REF] Glasby | Urban structures as marine habitats[END_REF]Davis et al., 2002, Bacchiocchi and[START_REF] Airoldi | The effects of sedimentation on rocky coast assemblages[END_REF][START_REF] Bulleri | Artificial marine structures facilitate the spread of a nonindigenous green alga, Codium fragile ssp tomentosoides, in the North Adriatic Sea[END_REF][START_REF] Glasby | Nonindigenous biota on artificial structures: could habitat creation facilitate biological invasions?[END_REF], with important implications for biodiversity at local and regional scales [START_REF] Bulleri | The introduction of artificial structures on marine soft-and hard-bottoms: ecological implications of epibiota[END_REF]. Some types of hard coastal-defence structures, such as breakwaters, not only introduce hard substrata in otherwise sandy-dominated bottoms, but can also provide sheltered habitats along wave-exposed coasts. Wave-exposure is an important determinant of the structure of benthic assemblages [START_REF] Denny | Predicting physical disturbance: mechanistic approaches to the study of survivorship on wave-swept shores[END_REF]. Thus, the availability of sheltered habitats can allow the development of assemblages very different from those occurring on nearby wave-exposed habitats and create, directly or indirectly, opportunities for the establishment, reproduction and spread of nonindigenous species [START_REF] Bulleri | Artificial marine structures facilitate the spread of a nonindigenous green alga, Codium fragile ssp tomentosoides, in the North Adriatic Sea[END_REF][START_REF] Bulleri | Development of reproductive structures in the introduced green alga, Codium fragile ssp. tomentosoides, in the northern Adriatic Sea[END_REF].

At odds with the ubiquitous presence of coastal-defence schemes and with the pressing need to improve our understanding of their impacts on natural assemblages at a variety of spatial scales and of their role as surrogates of natural rocky habitats [START_REF] Bulleri | Is it time for urban ecology to include the marine realm?[END_REF], very few studies [START_REF] Bacchiocchi | Distribution and dynamics of epibiota on hard structures for coastal protection[END_REF][START_REF] Moschella | Low-crested coastal defence structures as artificial habitats for marine life: using ecological criteria in design[END_REF][START_REF] Pinn | The assemblages on groynes in relation to substratum age, aspect and microhabitat[END_REF][START_REF] Gacia | Low crested coastal defence structures on the Catalan coast of the Mediterranean Sea: how they compare with natural rocky shores[END_REF] have investigated patterns of distribution of epibiota they support and, in particular, assessed differences between wave-exposed and wave-sheltered habitats [START_REF] Davis | Artificial armored shorelines: sites for opencoast species in a southern California Bay[END_REF].

This study, by assessing patterns of abundance and distribution of epibiota on breakwaters along sandy stretches of coastline in Tuscany (NW Mediterranean), is an attempt in that direction. Specifically, we tested the hypotheses that low-shore assemblages would differ between the landward (wave-sheltered) and the seaward (wave-exposed) side of breakwaters. The generality of patterns was evaluated by making these comparisons on four occasions, over a period of 15 months. Furthermore, because knowledge about the stability of assemblages, either in space or time, is important to predict the susceptibility of assemblages to the establishment of exotic species [START_REF] Davis | Fluctuating resources in plant communities; a general theory of invasibility[END_REF], we tested the hypothesis that temporal and spatial variation of the assemblages would be different between the landward and the seaward side of breakwaters.

Methods

Study site

The study was done on hard coastal-defence structures at Marina di Pisa, about 14 km north of Livorno, Tuscany (Italy), in the North Western Mediterranean (Fig. 1). The coast is sandy, exposed to westerly winds and extends from the mouth of the Arno River for about 15 km south.

During the mid Sixties', several offshore breakwaters running parallel to the coast were deployed to prevent the erosion of sandy shores. These are about 200 m in length and are made of granite blocks of irregular shape (major axis varying between 2 and 3 m). Breakwaters extend about 2-3 m below the Mean Low Water Level (hereafter MLWL), are between 30 and 50 m offshore and separated by narrow channels, about 5 m in width (Fig. 1). The spatial arrangement of breakwaters creates a calm internal water body on the landward side, characterized by soft bottoms and highly frequented by people from May to October (authors' personal observation).

Sampling Design

In October 2006, we started a preliminary survey to assess patterns of distribution and abundance of algae and invertebrates present on breakwaters. Two breakwaters were haphazardly selected (100s of m apart) along 2.5 km of coastline. Both the landward (wave-sheltered) and the seaward (wave-exposed) sides of breakwaters were sampled at low-shore levels (between 0 and 0.3 m below the MLWL). Three areas (about 5 m long and 10s of m apart) were then randomly selected within each side of each breakwater. The cover of sessile organisms and bare rock was quantified visually in five randomly placed 20 x 20 cm quadrats in each area, subdivided into 25 sub-quadrats (4 x 4 cm). A score from 0 to 4 % was given to each taxon in each sub-quadrat and the percentage cover was obtained by summing over the entire set of sub-quadrats [START_REF] Dethier | Visual versus random-point percent cover estimations: 'Objective' is not always better[END_REF]. When possible, organisms were identified in the field to species or genus; when it was not possible, taxa were lumped into morphological groups [START_REF] Steneck | A functional group approach to the structure of algaldominated communities[END_REF].

In December 2006, a greater number of breakwaters was included in our sampling design in order to improve our ability to estimate variation in assemblages at this spatial scale (i.e. among breakwaters). Thus, four breakwaters were randomly selected from a larger number available and, on each of these, two areas (about 5 m long and 10s of m apart) were randomly identified on both the landward and the seaward side. Different areas were selected for each time of sampling (December 2006, May andOctober 2007), to ensure data independence [START_REF] Underwood | Experiments in Ecology. Their logical design and interpretation using analysis of variance[END_REF]. Within each area, low-shore assemblages were sampled in five replicate quadrats with the same procedures previously described.

Data analysis

To test for differences in the structure of assemblages between exposures, data from the first time of sampling (October 2006) were analysed by permutational multivariate analysis of variance (PERMANOVA, [START_REF] Anderson | A new method for non-parametric multivariate analysis of variance[END_REF]. The analysis included the following factors: (i) Exposure, fixed, 2 levels (landward and seaward side of breakwaters); (ii) Breakwater, random, 2 levels, crossed with Exposure; (iii) Area, random, 3 levels, nested within the interaction of the other factors.

Student's t-test was used for multivariate pair-wise a posteriori comparisons.

Multivariate patterns were displayed graphically by plotting the centroids of areas in a nMDS (non-metric Multidimensional Scaling) based on Euclidean distances. In order to obtain centroids of each area for each date of sampling, we first calculated a dissimilarity matrix based on the Bray-Curtis index among all the observations. Because Bray-Curtis is a semi-metric index, centroids cannot be obtained simply as arithmetic averages of these dissimilarities [START_REF] Anderson | A new method for non-parametric multivariate analysis of variance[END_REF]). Thus, we first calculated principal coordinates from the Bray-Curtis dissimilarity matrix. This places the observations into a Euclidean space without altering the Bray-Curtis measure: i.e. the distance between any pair of observations based on the principal coordinates is equivalent to the dissimilarity between those observations obtained from the original variables. Centroids were then obtained as arithmetic averages of the principal coordinates over the 5 replicates of each area in each date of sampling.

The abundance of common taxa were analysed with analysis of variance (ANOVA), following the same design used for the PERMANOVA. Prior to analyses, Cochran's C test was used to assess the assumption of homogeneity of variances and data were Ln(x+1)-transformed, if necessary [START_REF] Underwood | Experiments in Ecology. Their logical design and interpretation using analysis of variance[END_REF]. Student -Newman -Keuls (SNK) tests were used for pairwise a posteriori comparisons of means. Data from the subsequent three times of sampling were analysed with the same multivariate and univariate techniques, but with a slightly different design. This included the following factors:

(i) Date, random, 3 levels; (ii) Breakwater, random, 4 levels, crossed with Date; (iii) Exposure, fixed, 2 levels (landward and seaward side of breakwaters), crossed with both Date and Breakwater;

(iv) Area, random, 2 levels, nested within the interaction of the other factors. Furthermore, SIMPER analysis [START_REF] Clarke | Non-parametric multivariate analyses of changes in community structure[END_REF] was used to identify those taxa that most contributed to Bray-Curtis dissimilarities between the seaward and landward sides of breakwaters at each date of sampling.

Univariate methods for partitioning variation among factors have been successfully extended to multivariate data, providing new insights into assemblage heterogeneity in marine environments [START_REF] Terlizzi | Scales of spatial variation in Mediterranean subtidal sessile assemblages at different depths[END_REF]). Estimates of temporal variation in structure of assemblages were obtained by running a 2-way PERMANOVA including the factors Time (random, 3 levels) and Area (random, 2 levels and nested in Time), separately for each combination of breakwater and wave-exposure. The multivariate pseudo-variance components for each term in the model were then calculated using multivariate analogues to the univariate ANOVA estimators (e.g. [START_REF] Searle | Variance components[END_REF][START_REF] Terlizzi | Scales of spatial variation in Mediterranean subtidal sessile assemblages at different depths[END_REF]). Observed mean squares were equated to the expected mean square derived from the linear model of the analysis [START_REF] Underwood | Experiments in Ecology. Their logical design and interpretation using analysis of variance[END_REF]. This procedure generated a total of 8 replicate measures of temporal variation, 4 for each of the seaward and the landward sides of breakwaters. Estimates of spatial variation between areas were obtained using a similar procedure; for each date of sampling, a 1-way PERMANOVA was performed for each breakwater and wave-exposure, providing 4 measures of spatial variation between areas for each side of breakwaters. Occasionally, negative estimates of pseudo variance were obtained. In these cases, variances were set to zero, assuming that they were sample underestimates of small or zero variances [START_REF] Searle | Variance components[END_REF][START_REF] Underwood | Spatial patterns of variability in density of intertidal populations[END_REF].

Results

The multivariate analysis on data collected in October 2006 showed significant differences in the structure of assemblages between the landward and the seaward side of breakwaters (Table 1).

In the nMDS, symbols representing assemblages on the landward side were segregated from those representing assemblages on the seaward side of breakwaters (Fig. 2).

Although there was a marked trend for the cover of Mytilus galloprovincialis and Corallina elongata, to be greater on the seaward than on the landward side of breakwaters, the ANOVAs failed to detect significant effects of wave-exposure, likely due to the lack of power of the relevant test and to heterogeneity of variances (Table 2, Fig. 3A and B respectively). In contrast, the covers of filamentous algae and Caulerpa racemosa and the amount of bare rock were significantly greater on the landward than on the seaward side of breakwaters (Table 2, Fig. 3C, D and E respectively), while the cover of encrusting corallines did not vary according to wave-exposure (Table 2, Fig. 3F).

All analysed taxa, excluding C. racemosa and C. elongata, showed significant variability among areas. The number of taxa was generally small and did not differ significantly between exposures (Table 2, Fig. 5).

For the three subsequent dates of sampling, when more breakwaters were included in the survey, multivariate analyses showed significant differences in the structure of assemblages between landward and seaward side of breakwaters that were consistent through time (t-tests for the interaction Date x Exposure Table 3, Fig. 4). SIMPER analyses showed that these differences were consistently caused by filamentous algae, C. elongata and M. galloprovincialis, explaining together 79%, 75% and 70% of the differences, in date 1, 2 and 3, respectively. Interestingly, in date 3 (October 2007), C. racemosa also contributed to differences between the landward and seaward side of breakwaters, explaining 11% of the differences. 3, Fig. 3D). The cover of encrusting corallines, despite largely varying through time, was greater on the seaward than on the landward side of breakwaters (Table 3, Fig. 3F), while the opposite pattern was detected for the amount of free space (Table 3, Fig. 3E). No significant differences were detected between sides of breakwaters in the number of taxa (Table 3, Fig. 5).

Visual inspection of graphs indicates that temporal variation in the structure of assemblages was greater in assemblages on the seaward side than those on the landward side of breakwaters (Fig. 6A). In contrast, variation within breakwaters, at the scale of 10s of m, did not vary according to wave-exposure (Fig. 6B).

Discussion

Low-shore assemblages on artificial structures, either on landward or seaward sides of breakwaters, were characterized by a low number of taxa, in accordance with patterns documented on breakwaters along other sedimentary coasts in the Mediterranean and in the North-east Atlantic [START_REF] Bacchiocchi | Distribution and dynamics of epibiota on hard structures for coastal protection[END_REF][START_REF] Moschella | Low-crested coastal defence structures as artificial habitats for marine life: using ecological criteria in design[END_REF][START_REF] Pinn | The assemblages on groynes in relation to substratum age, aspect and microhabitat[END_REF][START_REF] Gacia | Low crested coastal defence structures on the Catalan coast of the Mediterranean Sea: how they compare with natural rocky shores[END_REF]. These patterns are not surprising given that artificial structures, being generally located in urban or industrial areas, are commonly subjected to poor environmental conditions (e.g. release of pollutants and nutrients, high rates of sedimentation), which are known to affect negatively macroalgal diversity [START_REF] Krause-Jensen | Total and opportunistic algal cover in relation to environmental variables[END_REF].

At low-shore levels, the artificial habitats we investigated supported a smaller number of taxa (mean ± SE = 4.45 ± 0.09; n = 240) than nearby rocky shores, located about 15 km south (11.16 ± 0.18; n =192; Benedetti-Cecchi et al., unpublished data). Assemblages on artificial structures represented a subset of the regional pool of hard-bottom species, as all the species they supported were common on local rocky shores. This would indicate that a relatively large number of species is not able to establish viable populations on these surfaces. Severe sand scouring during swells may reduce the number of species able to recruit and/or survive on these structures. Alternatively, lack of supply of larvae, spores or propagules for colonization could explain the low species diversity. This is, however, unlikely to occur, given the relatively short distance between this site and natural rocky shores or reefs. Our results, in accordance with those of other studies [START_REF] Bacchiocchi | Distribution and dynamics of epibiota on hard structures for coastal protection[END_REF][START_REF] Moschella | Low-crested coastal defence structures as artificial habitats for marine life: using ecological criteria in design[END_REF][START_REF] Pinn | The assemblages on groynes in relation to substratum age, aspect and microhabitat[END_REF][START_REF] Gacia | Low crested coastal defence structures on the Catalan coast of the Mediterranean Sea: how they compare with natural rocky shores[END_REF] would suggest that these artificial habitats do not represent surrogate habitats for many rocky-bottom species, allowing few "opportunistic" organisms to achieve dominance.

Variation between the landward and the seaward side of breakwaters was mainly caused by four taxa, with encrusting corallines, Corallina elongata and Mytilus galloprovincialis more abundant on the seaward side of breakwaters and filamentous algae on the landward side. Other studies have reported similar patterns of distribution of organisms around breakwaters [START_REF] Bacchiocchi | Distribution and dynamics of epibiota on hard structures for coastal protection[END_REF][START_REF] Bulleri | Artificial marine structures facilitate the spread of a nonindigenous green alga, Codium fragile ssp tomentosoides, in the North Adriatic Sea[END_REF][START_REF] Moschella | Low-crested coastal defence structures as artificial habitats for marine life: using ecological criteria in design[END_REF][START_REF] Gacia | Low crested coastal defence structures on the Catalan coast of the Mediterranean Sea: how they compare with natural rocky shores[END_REF]. Mussels and barnacles have been repeatedly found to be more abundant on wave-exposed sides of artificial structures [START_REF] Bacchiocchi | Distribution and dynamics of epibiota on hard structures for coastal protection[END_REF][START_REF] Moschella | Low-crested coastal defence structures as artificial habitats for marine life: using ecological criteria in design[END_REF][START_REF] Gacia | Low crested coastal defence structures on the Catalan coast of the Mediterranean Sea: how they compare with natural rocky shores[END_REF]. At our study sites, the landward side of breakwaters was completely free from mussels. This could be the result of poor circulation of water reducing the supply of food or enhancing the deposition of fine sediments on the surface of blocks. In accordance with [START_REF] Gacia | Low crested coastal defence structures on the Catalan coast of the Mediterranean Sea: how they compare with natural rocky shores[END_REF], much of the free space on the landward side of breakwaters was indeed covered by a thick layer of sediment (F. Bulleri, personal observation). High sedimentation rates can cause smothering of filter-feeders like barnacles and mussels, favouring the dominance by filamentous or turf-forming algae [START_REF] Airoldi | The effects of sedimentation on rocky coast assemblages[END_REF]. Interestingly, throughout the duration of the study there was a trend for M. galloprovincialis to progressively monopolize space at the expense of C. elongata. This could indicate an ongoing process of recovery of space by mussel beds after a particularly severe storm disrupted their integrity or, alternatively, could reflect a successful event of recruitment. In contrast to patterns documented by Airoldi et al. (2005b) in the northern Adriatic Sea, recreational harvesting of mussels is not a common practice at our study site.

The main space-occupiers on the seaward side of breakwaters, M. galloprovincialis and C. elongata, are commonly found on nearby wave-exposed natural rocky shores (Livorno, about 15 km south), but never at such high values of percentage cover [START_REF] Menconi | Spatial and temporal variability in the distribution of algae and invertebrates on rocky shores in the northwest Mediterranean[END_REF][START_REF] Benedetti-Cecchi | Variability in abundance of algae and invertebrates at different spatial scales on rocky sea shores[END_REF]. Either physical features of artificial habitats (e.g. size and topography of the blocks and their spatial arrangement) or water quality could have contributed to determine their dominance at our study site. The proximity of breakwaters to the mouth of the Arno River might have created highly suitable conditions for the recruitment and growth of mussels (i.e. supply of suspended organic matter). Great similarity between assemblages in wave-exposed habitats on these breakwaters and those on a rocky shore influenced by the Magra River, about 50 km north, (i.e. dominance of mussels; [START_REF] Benedetti-Cecchi | The interplay of physical and biological factors in maintaining mid-shore and low-shore assemblages on rocky coasts in the northwest Mediterranean[END_REF] would suggest a strong influence of inputs of freshwater at our study site. The relatively large cover of Corallina elongata, a stenohaline species [START_REF] Doty | The distribution of marine algae into estuarine waters[END_REF], suggests however that factors other than the proximity to a river contribute to regulate the structure of low-shore assemblages on these structures.

Filamentous algae were present on the seaward side of breakwaters, but exclusively as epiphytes of mussels and Corallina elongata and with temporally variable percentage cover, greater in summer/autumn and close to zero in winter. In contrast, filamentous algae dominated wavesheltered habitats consistently through time. Here, it is worth stressing that, while mosaics of patches occupied by mussels or C. elongata have been documented in the region [START_REF] Benedetti-Cecchi | The interplay of physical and biological factors in maintaining mid-shore and low-shore assemblages on rocky coasts in the northwest Mediterranean[END_REF], persistent domination of space by filamentous forms has not been previously reported on open coasts. The provision of sheltered habitats may therefore generate physical conditions leading to the monopolization of space at low-shore levels by opportunistic forms and, ultimately, to the persistence of assemblages very different from those observed on natural rocky substrata in the region. If assemblages on the seaward side of breakwaters resemble, to some extent, those found on rocky shores in the region that are subjected to a similar influence of freshwater inputs, those on the landward side are to be considered habitats of their own, with artificial patterns and processes [START_REF] Bulleri | The introduction of artificial structures on marine soft-and hard-bottoms: ecological implications of epibiota[END_REF].

Assemblages on landward and seaward sides of breakwaters also differed in terms of temporal stability. Mechanical disturbance by waves is likely to vary greatly in intensity and frequency between exposed and sheltered habitats. Availability of bare space was greater on the landward than on the seaward side of breakwaters, consistently through time, suggesting that factors other than the wave-generated removal of biomass produced the observed patterns. Great availability of empty space has been previously documented on rip-rap seawalls built in wave-sheltered areas of San Diego Bay, in southern California [START_REF] Davis | Artificial armored shorelines: sites for opencoast species in a southern California Bay[END_REF]. Due to poor water exchange, conditions on the landward side of breakwaters could be sufficiently harsh (e.g. high rates of deposition of sediments) to allow the persistence of few species, with little temporal variation in their covers.

Indeed one single functional group, the filamentous algae, and, to a lesser extent, one exotic species, C. racemosa, were able to hold space consistently through time. Both of these are highly tolerant to sedimentation [START_REF] Airoldi | The effects of sedimentation on rocky coast assemblages[END_REF][START_REF] Piazzi | Sediment as a constituent of Mediterranean benthic communities dominated by Caulerpa racemosa var. cylindracea[END_REF]. Filamentous forms generally formed dense mats, making identification to species or genus unfeasible in the field. The same species were, however, found at different times of the year, suggesting low species turnover within this functional group (authors' personal observation). Variation between assemblages on the landward and the seaward side of breakwaters did not, in contrast, differ at the scale of 10s of m (i.e. between areas). This would indicate that although the factors influencing the distribution of organisms in these contrasting habitats are likely to differ (i.e. wave-action and sedimentation, in exposed and sheltered habitats, respectively), they operated consistently within breakwaters. Dominance by filamentous algae, associated with large availability of unoccupied space, on the landward side of breakwaters, may acquire great importance when considering the ecology of invasive species. In this study, we found a high cover of the introduced macroalga Caulerpa racemosa on the landward side of breakwaters. Although this alga is able to colonize a variety of natural substrata [START_REF] Piazzi | Invasion of Caulerpa racemosa var. cylindracea (Caulerpales, Chlorophyta) in the Mediterranean Sea: an assessment of the early stages of spread[END_REF]), here, it was not found on sandy bottoms sheltered by breakwaters. C. racemosa can tolerate high rates of sedimentation [START_REF] Piazzi | Sediment as a constituent of Mediterranean benthic communities dominated by Caulerpa racemosa var. cylindracea[END_REF] and its colonization and spread are facilitated by the presence of algal turfs [START_REF] Ceccherelli | Spread of introduced Caulerpa species in macroalgal habitats[END_REF]Bulleri and Benedetti-Cecchi, in press). Hence, domination by algal turfs in wave-sheltered habitats could have favoured the establishment of this invader. Importantly, this species is commonly found on exposed coasts and its distribution does not seem to be regulated by wave-exposure (Bulleri and Benedetti-Cecchi, in press). Hence, lack of this species on the seaward side of breakwaters may suggest that organisms that occupied space on these surfaces were somehow effective in preventing its establishment. Recent experimental work has shown that the spread and growth of C. racemosa is enhanced on complex surfaces (Bulleri and Benedetti-Cecchi, in press); according to these findings, mussel beds would represent suitable surfaces for C. racemosa, favouring the attachment of its prostrate stolons. Mussels and organisms intimately associated with them, such as epiphytes or those interstitially, are susceptible to dislodgement by waves. This mechanism is, however, unlikely to explain the complete absence of C. racemosa in wave-exposed artificial habitats.

Greater stability of assemblages on the landward side of breakwater could foster the persistence of C. racemosa. In the Mediterranean Sea, this alga shows strong seasonal fluctuations and, on the closest shallow natural rocky substrata (3-6 m in depth), it virtually disappears from November to March, to grow back in early spring and peak in abundance in September-October (Bulleri and Benedetti-Cecchi in press). Here, C. racemosa showed a similar trend, but its cover was still very high in early winter, a time of the year at which it has generally undergone a considerable decrease (Bulleri and Benedetti-Cecchi in press). Thus, wave-sheltered artificial habitats may not only provide suitable hard substrata for this species within an otherwise sandy area, but would also enhance its persistence through time. This may, in turn, enable this species to generate more biomass and produce more and bigger fragments, enhancing its long distance dispersal.

Although its cover did not reach values worth of formal analysis, a second non-indigenous species, the green alga Codium fragile spp. tomentosoides, was recorded on the landward side of breakwaters in December 2007. The results of this study, documenting the lack of C. fragile ssp.

tomentosoides on the seaward side of breakwaters, confirm that wave-sheltered artificial habitats represent ideal habitats for this species [START_REF] Bulleri | Artificial marine structures facilitate the spread of a nonindigenous green alga, Codium fragile ssp tomentosoides, in the North Adriatic Sea[END_REF][START_REF] Bulleri | Development of reproductive structures in the introduced green alga, Codium fragile ssp. tomentosoides, in the northern Adriatic Sea[END_REF]. Since , in contrast to C. racemosa, the recruitment of parthenogenetic zygotes C. fragile ssp. tomentosoides is enhanced on primary substrata [START_REF] Bulleri | Artificial marine structures facilitate the spread of a nonindigenous green alga, Codium fragile ssp tomentosoides, in the North Adriatic Sea[END_REF], the large availability of unoccupied space on the landward side of the breakwaters at Marina di Pisa let us predict a further increase in the local abundance of this species.

In summary, the results of this study suggest that the deployment of hard coastal-defence structures, attracting a suite of rocky-bottom species, has the potential to alter patterns of abundance and distributions of species at local and regional scale [START_REF] Bulleri | The introduction of artificial structures on marine soft-and hard-bottoms: ecological implications of epibiota[END_REF]. These structures can function as "stepping stones" (sensu [START_REF] Glasby | Urban structures as marine habitats[END_REF], enabling species to disperse across areas lacking of suitable habitat, disrupting natural patterns of dispersion and influencing gene flow, with potential evolutionary consequences. In particular, the provision of wave-sheltered surfaces determines the occurrence of assemblages different from those that can be found on comparable natural hard substrata. These are characterized by low functional diversity (one morphological group of algae), large availability of free space and great temporal stability. These features can have positive implications for the establishment and spread of exotic species [START_REF] Davis | Fluctuating resources in plant communities; a general theory of invasibility[END_REF][START_REF] Stachowicz | Biodiversity, invasion resistance, and marine ecosystem function: reconciling pattern and process[END_REF]. This study, documenting the presence of two of the most widespread invasive seaweeds [START_REF] Williams | A global review of the distribution, taxonomy, and impacts of introduced seaweeds[END_REF], strengthen the results of previous works that have identified man-made structures as important corridors for the expansion of introduced species [START_REF] Glasby | Urban structures as marine habitats[END_REF][START_REF] Bulleri | Artificial marine structures facilitate the spread of a nonindigenous green alga, Codium fragile ssp tomentosoides, in the North Adriatic Sea[END_REF].

In this light, further research on the ecological impacts and ecological role of these structures, whose presence is predicted to increase due to climatic and socio-economic reasons, is urgently needed to plan strategies for the conservation of marine biodiversity and for the management of biological invasions. Our study suggests that the provision of sheltered habitats should be of particular concern when planning the deployment of hard coastal-defence structures. Alternative options in the design of these structures may minimize their impacts. For instance, deploying structures that allow higher circulation of water around landward surfaces would serve to improve physical conditions, enabling the development of assemblages more similar to those occurring on natural rocky coasts. Achieving these goals requires, however, the integration of ecological criteria into the decision-making process which is, at present, mostly based on social and economic objectives. Whenever the analysis showed a significant interaction, the F-ratio for the main effect and lower order interactions of relevant factors are not reported as these are not logically interpretable [START_REF] Underwood | Experiments in Ecology. Their logical design and interpretation using analysis of variance[END_REF]. When necessary, pooling procedures were applied according to [START_REF] Underwood | Experiments in Ecology. Their logical design and interpretation using analysis of variance[END_REF]. L = Landward; S = Seaward.

Legend fo Figures

Caulerpa

  Univariate analyses showed the dominance of M. galloprovincialis and C. elongata on the seaward side and, although to a different extent among sampling dates, that of filamentous algae on the landward side of breakwaters (SNK tests in Table 4, Fig. 3A, B and C). The cover of C. racemosa varied markedly among dates of sampling. The alga was absent in May 2007 while it was found exclusively on the landward side of breakwaters in December 2006 and in October 2007, with a greater cover at the latter time of sampling (Table

Fig. 1 .

 1 Fig. 1. Map of the study area, showing the strecth of shores where breakwaters have been deployed.

Fig. 2 .

 2 Fig. 2. nMDS plot comparing assemblages on different side of breakwaters during October 2006.

Fig. 3 .

 3 Fig. 3. Abundance of common taxa and bare rock at each date of sampling. Bars are the average

Fig. 4 .

 4 Fig. 4. nMDS plots comparing assemblages on different sides of breakwaters in December 2006

Fig. 6 .

 6 Fig. 6. Multivariate temporal (A) and spatial (B) variance illustrating differences between

Fig

  Fig. 2 Vaselli et al.

Table 1 .

 1 PERMANOVA testing for the effects of wave-exposure and breakwaters on the structure of low-shore assemblages in October 2006.

	Source of

Table 2 .

 2 ANOVAs testing for the effects of wave-exposure and breakwaters on the abundance of common taxa in October 2006. *, P <0.05; **, P <0.01; ***, P <0.001. When necessary, pooling procedures were applied according to[START_REF] Underwood | Experiments in Ecology. Their logical design and interpretation using analysis of variance[END_REF].

			Caulerpa	Filamentous	Corallina	Encrusting	Mytilus
			racemosa	algae		elongata	corallines	galloprovincialis
	Source of Variability d.f.	MS	F	MS	F	MS	F	MS	F	MS	F
	Exposure = E	1	8954.81 a 20.43**	286.97 410.64*	69292.01 50.47	0.71 4.71	17442.15 9.91
	Breakwater = B	1	432.01 0.99	33.47	1.53	1450.41 2.35	0.06 0.05	1760.41 2.22
	E x B	1	268.81 Eliminated	0.69	0.03	1372.81 2.23	0.15 0.12	1760.41 2.22
	Area(E x B)	8	438.21 1.90	21.89	7.39***	616.10 1.50	1.27 2.36*	792.28 3.59**
	Residual	48	230.10		2.96		411.45		0.53		220.49
	Cochran's Test		C=0.50**	C=0.27		C=0.49**	C=0.25	C=0.31*
	Transformation		None		Sqrt(x+1)	None		Ln(x+1)	None
	a Tested on Area(E x B)

a Tested on Area(E x B)

Table 3 .

 3 PERMANOVA testing for the effects of date of sampling, breakwater and wave-exposure on the structure of low-shore assemblages. Whenever the analysis showed a significant interaction, the pseudo-F for the main effect and lower order interactions of relevant factors are not reported as these are not logically interpretable[START_REF] Underwood | Experiments in Ecology. Their logical design and interpretation using analysis of variance[END_REF]).

	L = Landward; S = Seaward.
	Source of

Table 4 .

 4 ANOVAs testing for the effects of date of sampling, breakwater and wave-exposure on the abundance of common taxa. *, P <0.05; **, P <0.01; ***, P <0.001.

Acknowledgements

We sincerely thank I. Bertocci, E. Maggi and L. Tamburello for their help in the field. This study was partially supported by the Comune di Pisa and it is publication number MPS-xx of MARBEF Network of Excellence. Procedures used in the present study comply with Italian laws.