Andreia Cruz

Tânia Caetano

Satoru Suzuki

Sónia Mendo
email: smendo@bio.ua.pt.

a tributyltin (TBT) degrading bacterium isolated from an estuarine environment, Ria de Aveiro in Portugal

Organotin compounds are used in a variety of industrial processes therefore their subsequent discharge into the environment is widespread.

Bacteria play an important role in biogeochemical transformations acting as natural decontamination agents. Therefore, screening for tributyltin (TBT)-resistant and -degrading bacteria is relevant for the selection of isolates with decontamination ability of these polluted areas. With this purpose, fifty strains were isolated from sediment and water from Ria de Aveiro and their tolerance to TBT, up to 3 mM, was evaluated. Generally, occurrence of highly TBTresistant bacteria was observed, and Gram negative bacteria exhibited more tolerance to TBT than Gram positive bacteria. A memory response was observed when bacteria were progressively exposed to increasingly higher TBT concentrations. One isolate, Aeromonas veronii Av27, highly resistant to TBT (3 mM) uses this compound as carbon source and degrades it to less toxic compounds.

INTRODUCTION

Tributyltin (TBT) is an organotin compound broadly applied as fungicide, bactericide, pesticide, wood preservative, PVC stabilizer and component of antifouling paints. Although the production, use and export of TBT have been prohibited in developed countries since the 1990s by the International Maritime Organization (IMO), some countries have continued to utilize this agent until its complete prohibition, which was scheduled to go into effect in 2008 [START_REF] Rudel | Retrospective monitoring of organotin compounds in marine biota from 1985 to 1999: results from German environmental specimen bank[END_REF].

Consequently, TBT pollution has been recognized as a serious problem around the world. Ria de Aveiro is an important estuarine system on the north-west coast of Portugal. It is a shallow system that occupies an area of 43 km 2 at low tide and its bottom consists mainly of muddy sediments. The most important sources of organotin pollution are the port, dockyards and marinas, where TBT levels range from 9-42 ng TBT-Sn l -1 in the water and up to 88 ng TBT-Sn g -1 dry weight total sediment [START_REF] Barroso | Comparison of imposex and intersex development in four prosobranch species for TBT Monitoring of a southern European estuarine system (Ria de Aveiro, NW Portugal)[END_REF].

Organotins can be highly toxic to many eukaryotic and prokaryotic organisms and have also been identified as immune system inhibitors and endocrine disruptors in humans [START_REF] Gibbs | TBT-induced imposex in neogastropod snails: masculinization to mass extinction[END_REF][START_REF] Dubey | Biodegradation of tributyltins (organotins) by marine bacteria[END_REF][START_REF] Dubey | Microarray-mediated transcriptome analysis of the tributyltin (TBT)-resistant bacterium Pseudomonas aeruginosa 25W in the presence of TBT[END_REF]. In many aquatic organisms, its toxic effect is also recognized; the concentration of 1-2 ng l -1 (nanomolar level) causes growth suppression, immunosuppression, and imposex in higher animals [START_REF] White | Organotin compounds and their interaction with microorganisms[END_REF][START_REF] Hoch | Organotin compounds in the environment-an overview[END_REF].

Generally, trisubstituted organotins (TBT) are more toxic than di-(DBT) and monosubstituted (MBT) compounds, and seem to exert toxicity through their interaction with membrane lipids. It has been reported that organotin compounds are toxic to both Gram positive and Gram negative bacteria collected from sediment; nevertheless, the former showed increased sensitivity to triorganotins [START_REF] Mendo | Tributyltin and triphenyltin toxicity on benthic estuarine bacteria[END_REF].

Several mechanisms have been proposed that could be involved in TBT resistance in bacteria: (i) transformation into less toxic compounds (DBT and MBT) by abiotic and biotic mechanisms [START_REF] Clark | The fate of tributyltin in the aquatic environment[END_REF][START_REF] Dowson | Persistence and degradation pathways of tributyltin in freshwater and estuarine sediments[END_REF][START_REF] Pain | Characterization of organotin-resistant bacteria from Boston Harbor sediments[END_REF]; (ii) exclusion of the compound from the cell mediated by a multidrug efflux pump [START_REF] Jude | TbtABM, a multidrug efflux pump associated with tributyltin resistance in Pseudomonas stutzeri[END_REF]; (iii) degradation/ metabolic utilization as a carbon source [START_REF] Kawai | Degradation of tributyltin by a bacterial strain isolated from polluted river water[END_REF]; (iv) bioaccumulation into the cell without breakdown of the compound [START_REF] Fukagawa | Occurrence of tributyltin (TBT) and methyl mercury tolerant bacteria in natural sea water to which TBT was added[END_REF].

The catabolic activities of introduced microorganisms play an important role in bioremediation technologies for the restoration of polluted environments [START_REF] Heinaru | Biodegradation efficiency of functionally important populations selected for bioaugmentation in phenol-and oil-polluted area[END_REF]. In this context, a given bacterial population might perform the reverse reaction on tin compounds by demethylation or methylation of TBT into less toxic derivatives [START_REF] Jude | TbtABM, a multidrug efflux pump associated with tributyltin resistance in Pseudomonas stutzeri[END_REF]. Given the function played by bacteria in biogeochemical cycles, the identification and characterization of TBT pollutant-degrading bacterial strains is crucial. Knowledge of their physiology and genetics is fundamental for their future application as natural decontamination agents.

In the present study we aimed to isolate and identify bacteria resistant to TBT that could be used and developed as biosensors for TBT pollution. The ability to degrade TBT into DBT and MBT (less toxic compounds), and its use as carbon source, was evaluated on selected bacteria. One particular isolate, A. veronii Av27, fulfilled all the characteristics needed for its future application as a TBT sensor and also for the remediation of TBT polluted areas.

MATERIALS AND METHODS

Bacterial isolation and identification

Bacterial strains were isolated from sediment (November 2001) [START_REF] Mendo | Tributyltin and triphenyltin toxicity on benthic estuarine bacteria[END_REF] and water (Janeiro 2004) from Ria de Aveiro, a shallow estuarine system, near Aveiro (40º38'N, 8º41'W), in the north-west coast of Portugal. The TBT concentrations recorded were 39 ng Sn g -1 and 79 ng Sn l -1 in sediment and water, respectively [START_REF] Barroso | Comparison of imposex and intersex development in four prosobranch species for TBT Monitoring of a southern European estuarine system (Ria de Aveiro, NW Portugal)[END_REF].

Sediment cores (10 cm in diameter and 5 cm depth) were used to collect the sediment that was extruded into sterile flasks and processed immediately. Bacteria were isolated by the following steps: 10 g of sediment were mixed with 100 ml of sterile artificial seawater, for 30 min; serial dilutions, up to 10 -10 , were plated on Tryptic Soy Agar (TSA) (Merck, Germany), Plate Count Agar (PCA) (Merck, Germany) and Marine Broth Agar (MBA) (Difco, USA); Petri dishes were incubated overnight at 20 ºC.

Water samples were collected from overlaying water (20 cm) using an immersed sterile bottle during the flood tide. Samples were transported to the laboratory and processed within 4 h. Bacteria were isolated by following steps: serial dilutions up to 10 -6 in sterile artificial seawater were spread in MBA, TSA and PCA; plates were incubated overnight at 20 ºC.

Morphologically different colonies were picked from the plates and re-streaked on the same medium to ensure purity. Fifty isolates were selected and identified using Biolog Microlog 3 4.01 C, Vitek ® and API ® systems (bioMérieux, Hazelwood, MO, USA). Whenever necessary, 16S rDNA sequence was determined after PCR amplification of the corresponding amplicon [START_REF] Muyzer | Profiling of complex microbial populations by denaturing gradient gel electrophoresis analysis of polymerase chain reactionamplified genes coding for 16 rRNA[END_REF]. Nucleotide amino acid sequences were analysed using BioEdit Sequence Alignment Editor (Isis Pharmaceuticals) and the BLAST program available from the National Institute for Biotechnology Information server (http://www.blast.genome.ad.jp). Strains were maintained in

TBT sensitivity

Gradient plates were prepared with TBTCl (97 %) (Fluka, Switzerland) incorporated in TSA medium, with concentrations ranging from 0 to 3 mM (subsequently the compound will simply be referred to as TBT). TBT stock solution was prepared in ethanol and stored in the dark at 4 ºC.

Each plate, with the dimensions of 11x11 cm, consisted of a first layer of TSA (50 ml) containing 3 mM of TBT that was left to stand with a slight inclination; after the agar has solidified, 50 ml of TSA were layered on the top of the first layer (adapted from [START_REF] Metzenberg | Disruption of essential genes in Neurospora by RIP[END_REF]. Bacteria to be tested were grown on TSB at 250 rpm (Sanyo incubator shaker) until they reached an optical density (OD 600 nm ) of 1 (approximately after 10 h of incubation) at 26 ºC. A temperature of 26 ºC was chosen, as previous experiments showed that this temperature allowed a faster growth of the different bacteria. Each point of the gradient griddle was inoculated with 2 µl of culture. The tests were repeated in triplicate for each bacterial strain in different gradient plates. Isolates were scored as resistant if they can grow along the gradient griddle. In order to confirm the results obtained by this procedure, bacteria were then grown on TSA plates containing different nominal concentration of TBT (0, 0.025, 0.05, 0.075, 1, 1.5, 2, 3 mM). Plates were incubated at 26 ºC for 20 h in the dark.

Two isolates, one very resistant (Aeromonas veronii -Av27) and one sensitive (Shewanella algae/putrefaciens -Av35) were selected to compare their growth behaviour in liquid medium containing different TBT concentrations (0, 0.5, 1.5 and 3 mM). Growth occurred in TSB at 250 rpm, 26 ºC for 72 h, and was recorded as a change of optical density at A 600 nm .

Memory response to progressive TBT exposure

Bacteria that have the ability to grow on 0.05 mM of TBT were successively transferred to TSA plates containing increasingly higher TBT concentrations (1, 1.5 and 2 mM). Transfer to a new media occurred after 48, 72 and 96 h, as long as growth was observed. Plates were incubated at 26 ºC in the dark.

Antibiotic and heavy-metal sensitivities

Antibiotics to be tested were selected according to the Clinical Laboratory Standard Institute (CLSI) (Clinical Laboratory Standard Institute, 2005). The tests were performed by the Vitek ® system (bioMérieux, Hazelwood, MO, USA). Those included: amikacin ( 2 g ml -1 ), aztreonam ( 8 g ml -1 ), cefepime ( 4 g ml -1 ), ceftazidime ( 8 g ml -1 ), ciprofloxacin ( 0.5 g ml -1 ), gentamicin ( 0.5 g ml -1 ), imipenem ( 4 g ml -1 ), netilmicin ( 4 g ml -1 ), pefloxacin ( 1 g ml -1 ), piperacillin ( 8 g ml -1 ), piperacillin/tazobactam ( 8 g ml -1 ), ticarcillin ( 16 g ml -1 ), ticarcillin/clavulanic acid ( 16 g ml -1 ), tobramycin ( 0.5 g ml -1 ) and trimethoprim/sulfametoxazol ( 10 g ml -1 ). Results were interpreted according to the CLSI criteria (Clinical Laboratory Standard Institute, 2005).

Heavy-metal resistance was tested using gradient plates by the incorporation of the selected metal in TSA medium (as previously described in section 2.2). Metals and respective concentrations selected were based on previous determinations carried out in Ria de Aveiro [START_REF] Monterroso | Distribution and accumulation of metals (Cu, Cd, Zn and Pb) in sediments of a lagoon on the north-western coast of Portugal[END_REF]. Heavy-metal stock solutions were prepared in distilled water with the following molarities: Cu as CuSO 4 .5H 2 0, 0.2 M; Zn as ZnCl 2 , 0.1 M; Cd as CdCl 2 , 0.1 M; Hg as HgCl 2 , 0.1 M. Final concentrations of each of the heavy-metals were: 75 µg ml -1 of HgCl 2 , 3483 µg ml -1 of and 250 rpm (Sanyo incubator) until they reached an optical density (OD 600 nm ) of 1 (approximately 10 h of incubation). Each point of the gradient griddle was inoculated with 2 µl of culture. Tests were performed in triplicate for each bacterial strain in independent gradient plates.

Plates were incubated at 26 ºC for 20 h in dark conditions. Resistance profile was defined in function of growth observed along the gradient griddle.

TBT degradation by Aeromonas veronii Av27

To evaluate the ability of A. veronii Av27 to degrade TBT, a preliminary experiment was designed as follows: 100 ml of TSB medium containing 50 M of TBT were inoculated with 1 ml of an exponential phase culture of A. veronii (approximately 10 8 cells ml -1 ) and incubated for 24 h in an orbital shaker (Sanyo) at 26 ºC, 250 rpm, in the dark (to avoid photo degradation of TBT) (corresponding to sample II). A control sample, consisted of non-inoculated medium, put in the same conditions (corresponding to sample I).

After the incubation period, the bacterial culture was centrifuged at 8000 x g, for 10 min; cells were lyophilised and kept at -70 ºC until analysis (sample IIB); the resulting supernatant was also lyophilised and kept at -70 ºC until analysis (sample IIA). Control sample was subjected to lyophilisation prior to storage at -70 ºC, until analysis (sample I). Three independent replicates were performed for each experimental condition.

Organotin determinations were performed by gas chromatography/mass spectrometry (GC-MS).

The GC-MS method used is described in detail elsewhere [START_REF] Szpunar | Rapid speciation of butyltin compounds in sediments and biomaterials by CGC-MIPAES after microwave-assisted leaching/digestion[END_REF][START_REF] Quintela | The use of Nucella lapillus (L.) transplanted in cages to monitor tributyltin (TBT) pollution[END_REF]. In summary, three replicates of 0.1 g lyophilized cells/supernatant were digested (microwave assisted) with tetramethylammonium hydroxide. pH was adjusted to 5, sodium tetraethylborate and isooctane, containing tetrabutyltin as an internal standard, were added successively. After microwave treatment, the organic phase was recovered and analyzed by GC-MS. Recoveries of spiked organotins were, on average, 90% TBT, 98% DBT, 51% MBT. The determinations were performed at Servicios Xerais de Apoio á Investigación (Universidade da Coruña).

TBT utilization as carbon source and cell viability determination

Growth of A. veronii Av27 on TBT as the primary carbon source was evaluated as follows: 1 ml of an exponential phase culture of A. veronii (approximately 10 8 cells ml -1 ) was inoculated in Minimal Medium-MM -(KH 2 PO 4 1.0 g, MgSO 4 .7H 2 0 200.0 mg, FeSO 4 .7H 2 0 10.0 mg, CaCl 2 10.0 mg, vestigial elements (Mn, Mo, Cu, Co, Zn) as organic salts, 0.02 mg of each, NH 4 Cl 1.0 g, distilled water to 1l, pH 6.0) containing different concentrations of the TBT (0, 1.5, 3 mM) and incubated at 26 ºC and 250 rpm (Sanyo incubator shaker) for 72 h, in the dark.

TBT stock solution is prepared in ethanol; therefore, a control was prepared using the same volume of ethanol (30 ml l -1 ), to assure that the cell growth was supported by TBT. At 3 h intervals, 1 ml aliquots were withdrawn from the culture. Optical density was monitored at 600 nm (A 600 nm ), using a Lambda 16 Spectrophotometer (Cary 50B.O UV-visible spectrophotometer-Varian). Three independent replicates were performed for each experimental condition.

Bacterial cells viability was determined, using 1 ml of bacterial cultures (approximately 10 8 cells ml -1 ) grown in MM with (3 mM TBT) and without TBT. Three microliters of the dye mixture (propidium iodide and SYTO 9, LIVE/DEAD® BacLightTM of bacterial viability kit, Molecular Probes) was mixed with water to a final concentration of 30 µM propidium iodide and 5.01 µM Syto 9. The samples were incubated for 15 min in the dark at room temperature (RT), according to the manufacturer's instructions, filtered through 0.2 µm black polycarbonate membranes (Poretics) and mounted between a slide and cover slip with BacLight mounting oil (Molecular Probes). The green (intact membrane) and red (damaged membrane) were enumerated after 18-24 h using epifluorescence microscopy. Ten microscope fields or at least 200 cells were enumerated for each of 3 replicate samples.

RESULTS

Bacterial isolation, identification and TBT sensitivity

Fifty morphologically different strains were isolated from water and sediment samples collected at Ria de Aveiro. Forty six of those isolates were identified, whenever possible, to the species level (table 1). Four strains could not be identified. Growth along a TBT gradient, in solid medium showed that the sensitivity of the different strains to this compound was very heterogeneous (table 1). No consistency was observed at the species level, among the isolates in respect to their TBT resistance profile. At the genera level similar results were observed; for instance, in Bacillus (Av1, 2, 3 and 4) and Brevibacterium (Av5 and 6), where species belonging to same genera have a different behaviour in respect to TBT resistance. Among the isolates, 16 strains could grow on concentrations of up to 3 mM. Those strains belong to various genera. Some strains could not grow at all in the gradient plates. Nevertheless this may be due to some diffusion of the compound through the agar which means that TBT concentration never equals zero in the gradient plate and, therefore, growth is impaired even when the concentration is apparently close to zero. Thirteen strains were able to grow at intermediate TBT concentrations (0.025 mM -0.05 mM TBT) and, once again, significant divergent resistance profiles were observed in bacteria belonging to the same genera. Overall, the percentage of Gram positive bacteria isolated that were sensitive to TBT (no growth) was 43.5 % whilst Gram negative bacteria represented only 7.4 %. Although the samples were collected at the same site in Ria de Aveiro, no consistency was observed between bacteria TBT resistance profile and nature of the sample (water or sediment). Also, different resistance patterns were observed among bacteria isolated from water and from sediment.

Despite the fact that several bacterial isolates showed resistance to high TBT concentrations (3 mM), only one particular isolate, Av27, identified as Aeromonas veronii biovar sobria, was selected for further studies, because it possesses 4 plasmids, that might codify for TBT resistance, and also because its total DNA was more efficiently digested than the DNA of other resistant isolates, which would be a crucial step for the molecular studies.

Growth behaviour of Av27 on liquid medium with TBT

Figure 1 compares growth of Aeromonas veronii biovar sobria (Av27), in liquid culture media containing different TBT concentrations, with that of Shewanella algae/putrefaciens (Av35sensitive to TBT). It can be observed that strain Av27 can grow in the presence of high concentrations of TBT (up to 3 mM), while Av35 could not grow at the lowest concentration tested (0.5 mM).

'Memory response' to progressive TBT exposure

After the first screening for the isolation of TBT resistant strains, it was observed that only two Gram positive and seven Gram negative isolates could grow in 2 mM of TBT (table 1). However we noticed that, if those isolates that could not grow up to this concentration were successively transferred to increasingly higher TBT concentrations (up to 2 mM), the number of isolates increased to five Gram positive isolates and 14 Gram negative isolates. These results showed that, in vitro, bacteria can adapt to higher TBT concentrations, suggesting the existence of a 'memory response' to previous exposure.

Antibiotic and heavy-metal sensitivity of isolate Av27

A. veronii Av27 showed to be sensitive (no growth observed) to all the antibiotics tested (table 2).

Considering heavy-metal resistance pattern, A. veronii Av27 can grow up to 10 g ml -1 of HgCl 2 , 498 g ml -1 of CuSO 4 .5H 2 O, 86 g ml -1 of ZnCl 2 and up to 56000 g ml -1 of CdCl 2 (table 2).

TBT degradation by Aeromonas veronii Av27

Figure 2 represents organotin degradation determined when A. veronii Av27 was grown in TSB over a period of 24 h. Likewise table 3 shows the GC-MS values of organotin analysis. Sample I corresponds to non-inoculated culture media that was kept in the same conditions of incubation as the inoculated sample (sample II). It can be seen that no TBT degradation occurred under these conditions and mass spectrometry values for DBT and MBT were below the detection limit (table 3). Therefore, it can be assumed that 100 % of TBT remained in these samples. Samples IIA and IIB correspond to the culture supernatant of a 24 h growth of A. veronii and to the pelleted cells, respectively. In sample IIA, DBT was detected (120 µg ml -1 ) and TBT and MBT values were below the detection limit, suggesting TBT degradation has occurred and that the resulting compounds of degradation are extruded from the cell and released to the culture media.

Sample IIB reveals the amount of organotins that are present in the cells. The values correspond to about 89 % of TBT (0.32 µg ml -1 ), 10 % DBT (0.037µg ml -1 ), and only 1 % to MBT (0.0042 µg ml -1 ). These results suggest that either TBT remains adhered to the cells or part of it, that is degraded, remains in the cells.

TBT a carbon source for Aeromonas veronii Av27

Figure 3 shows the growth of A. veronii in appropriate culture medium in the presence of increasingly higher concentrations of TBT (0, 1.5, 3 mM). It can be observed that at a higher TBT concentration is lower (1.5 mM), lower optical density values are observed (A 600 nm of 0.5) at the same time interval. Similarly, when a control growth curve is done in the presence of ethanol, the TBT solvent, bacteria growth was slower (A 600nm = 0.3), suggesting that in fact, A. veronii uses TBT as carbon source for metabolic activities. Also, when A. veronii is grown in MM the total cell counts, as revealed by microscopy, were lower (120 x 10 6 ml -1 ) when compared to the total cell number observed when bacteria is growing in the presence of TBT (160 x 10 6 ml -1 ). Although the differences observed were not significant it is possible to suggest that TBT stimulates an increment in the cell number.

Microscope observations showed that when bacteria grow in MM without TBT, cells are bigger and appear dispersed in the culture medium, nonetheless when bacteria grow in presence of TBT, the cells look smaller and appear to aggregate in the culture medium.

Total viable counts showed that when A. veronii was grown in MM (without TBT), the number of viable cells is lower (22 %) than when bacteria was grown in MM with 3 mM TBT (45 %).

DISCUSSION

In the present study the gradient plate assay proved to be very useful method to detect different TBT resistance patterns among the bacterial isolates collected in Ria de Aveiro, allowing for the selection of a wide variety of cultivable bacterial genera exhibiting a TBT resistant phenotype.

Given the important role played by bacteria in biogeochemical processes, these bacteria consortia play an important role in estuarine systems.

A significant difference was observed between isolates belonging to the same genera. For instance, in the genera Bacillus, two strains Bacillus cereus (Av2) and Bacillus firmus (Av3) showed different tolerance to TBT (Table 1), despite the fact that they were both isolated from sediment samples. Similar observations were reported by Jude and co-workers (1996) in Pseudomonas species, where bacteria isolated from Arcachon harbor showed different TBT tolerance ranging from 0.07 and 3 mM.

In the present study, no significant correlation was observed between strains isolated from the more polluted area (water) and strains collected at the less polluted area (sediment). These findings were also observed by other researchers, suggesting that other factors rather than the presence of TBT select for a resistant population [START_REF] Wuertz | Tributyltin resistant bacteria from estuarine and freshwater sediments[END_REF][START_REF] Jude | Evaluation of tributyltin resistance in marine sediment bacteria[END_REF]. [START_REF] Ramamoorthy | Binding of mercuric and other heavy metal ions by microbial growth media[END_REF] suggested that medium-organometal interactions may affect metal toxicities. Nevertheless, despite the fact that TBT could act as a chelating agent sequestering carbon /nitrogen sources when a rich medium is used (e.g. TSB), this could not be observed in the present study, as bacteria retained the same TBT profile independently of the culture media used (MB, MBA, TSB, TSA). Furthermore, we found that despite the heterogeneity observed in the TBT resistance profile of Gram negative and Gram positive bacteria, we found that the former are generally more resistant to TBT. These findings were also observed by other authors and could be related with the different architecture of bacterial cell walls which, in Gram negatives, could promote the ability to immobilize metal ions [START_REF] Beveridge | Binding of metals to cell envelopes of Escherichia coli K-12[END_REF]. Nonetheless, five Gram positive strains were shown to be highly resistant to TBT. These findings suggest that, not only the external membrane but also cytoplasmic membrane and/or intracellular level mechanisms might be involved in the resistance, as suggested by [START_REF] White | Organotin compounds and their interaction with microorganisms[END_REF].

Successive transfer of the isolated bacteria to progressively higher TBT concentrations resulted in an increased resistance to TBT. These results are in accordance with those reported by Dahllöf and co-workers (2001), where in a marine sediment community, progressive exposure to higher TBT concentrations showed an increase in bacterial resistance. Also, [START_REF] Blanck | Pollution-induced community tolerance (PICT) in marine periphyton in a gradient of tributyltin (TBT) contamination[END_REF] reported that exposure to a toxicant for some time can increase tolerance to that particular toxicant. Therefore, according to our observations we proposed that this fact could reflect an inducible 'memory response' mechanism to TBT/toxicant exposure. Nonetheless, the relation between the 'memory response' to TBT concentration and the resistant bacteria on a given natural environment could not be made because other factors, for instance the presence of other toxic compounds, could influence the 'memory' mechanism response.

Due to the relevant aspects previously highlighted, A. veronii Av27 was selected for a series of studies which were conducted in the present work. This is the first report of an Aeromonas bacterium exhibiting high resistance to TBT. Until now, only members from Pseudomonas and Alteromonas genera were associated with TBT resistance. In those cases both strains were used to investigate the TBT resistance mechanism [START_REF] Suzuki | Tributyltin-resistant marine bacteria: a summary of recent work[END_REF][START_REF] Jude | TbtABM, a multidrug efflux pump associated with tributyltin resistance in Pseudomonas stutzeri[END_REF][START_REF] Roy | Tributyltin chloride-utilizing bacteria from marine ecosystem of west coast of India[END_REF], suggesting the existence of different mechanisms in those two genera.

A. veronii Av27 is sensitive to various antibiotics and did not show high tolerance to mercury. However, according to the resistant pattern observed in several environmental strains (e.g.

Pseudomonas, Proteus, Xanthomonas, Alteromonas, Aeromonas, and Enterobacteriaceae) studied by De and co-workers (2003), A. veronii, seems to be resistant to cadmium, copper and zinc, despite the fact that the values detected in Ria de Aveiro [START_REF] Monterroso | Distribution and accumulation of metals (Cu, Cd, Zn and Pb) in sediments of a lagoon on the north-western coast of Portugal[END_REF] are higher than those used by these authors. These results are quite intriguing, as we expected that A. veronii Av27 would also be resistant to several classes of antibiotics, considering that Ria de Aveiro is a reservoir of several resistance determinants (e.g. antibiotics and antibiotic gene cassettes) that are released from industrial and domestic effluents [START_REF] Henriques | Occurrence and diversity of integrons and beta-lactamase genes among ampicillin-resistant isolates from estuarine waters[END_REF]. Nevertheless, [START_REF] Jude | TbtABM, a multidrug efflux pump associated with tributyltin resistance in Pseudomonas stutzeri[END_REF] identified in Pseudomonas stutzeri, a multidrug efflux pump, belonging to the resistancenodulation-division (RND) family that conferred resistance to some toxic compounds and different classes of antibiotics. Degradation of TBT in DBT occurred under laboratory conditions by A. veronii Av27, in 24 h. Also, it seems that TBT is taken up by the cells, possibly, into siderophores structures, and it is gradually exported, as DBT, after degradation inside the cell. Despite the fact that debutylation of TBT results in less lipophilic compounds, due to cell activity, it appears that these compounds are still bound to the cell membrane lipids. Previously, [START_REF] Suzuki | Tributyltin-resistant marine bacteria: a summary of recent work[END_REF] reported that in a resistant strain, after 1 h of growth, the uptake of TBTCl was registered. However, during the incubation, a decrease of TBTCl uptake occurred probably due to efflux of the compound. These observations corroborate our results. Also, [START_REF] Roy | Tributyltin chloride-utilizing bacteria from marine ecosystem of west coast of India[END_REF] reported that a Pseudomonas sp. was able to use TBT as a source of carbon only at concentrations of up to 2 mM. In the present study, TBT was used as a source of carbon up to a concentration of 3 mM by A. veronii Av27, a concentration never reported before. Bacterial viability analysis showed that despite the higher using the available TBT, cells were smaller and appeared aggregated. This could be explained as an energy saving mechanism exhibited by the bacteria when exposed to TBT for long periods.

Despite their resistance to TBT bacteria cells are growing under stress conditions, due to the scarcity of nutrients. This was also observed and suggested by other authors for cells growing under stress conditions in the presence of other toxic compounds [START_REF] Dykhuizen | Metabolic flux and fitness[END_REF].

The results so far obtained suggest that A. veronii Av27 developed a physiological mechanism that gives the bacteria the ability to use TBT as a carbon source and to degrade it to less toxic compounds, conferring the advantage of resistance to high TBT concentrations.

CONCLUSION

Given its ability to degrade and utilize TBT as carbon source, A. veronii Av27 constitutes an important organism that could be used as a tool for genetic studies. These studies are already in progress with the aim of elucidating the mechanism behind TBT resistance at the genetic level. Furthermore, this strain might have the potential to be developed as a biosensor to detect TBT in the environment. Ultimately, A. veronii could be applied in the natural remediation of TBT contaminated areas.

FIGURES CAPTIONS

Figure 1. Growth of A. veronii Av27 in TSB medium with 0, 0.5, 1.5, 3 mM TBT and strain Av35 (sensitive to TBT) in TSB medium with of 0.5 mM TBT. Growth occurred at 26 ºC for 72 h and 250 rpm, and it was monitored as a change of absorbance at 600 nm. Standard deviations are indicated in the graph. Table 3. Organotin determinations by GC-MS. Sample I: TSB culture medium with 50 M of TBT (24 h incubation); Sample IIA: supernatant of TSB medium with 50 M TBT inoculated with A. veronii Av27 (24 h culture); Sample IIB: pellet of TSB medium with 50 M TBT inoculated with A. veronii Av27 (24 h culture). bdl: below detection limit.

Detection limits: monobutyltin (MBT) = 0.03 ng ml -1 ; dibutyltin (DBT) = 0.2 ng ml -1 ; tributyltin (TBT) = 0.07 ng ml -1 .

Organotin determinations by GC-MS MBT DBT TBT I bdl bdl 130 µg ml -1 IIA bdl 120 µg ml -1 bdl Samples IIB 0.0042 µg ml -1 0.037 µg ml -1 0.32 µg ml -1

Figure 2 .

 2 Figure 2. Percentage of organotins detected by mass spectrometry. Sample I: TSB culture

Figure 3 .

 3 Figure 3. Growth of A. veronii Av27 in Minimal Medium (MM) with different concentrations of

  Figure 1

ACKNOWLEDGMENTS

This work was supported by the following grants: POCI/MAR/56475/2004 (Fundação para a Ciência e a Tecnologia, Portugal), Fundação Oriente (Portugal), COE Program (Japan).


Maria José Brogueira

Maria Do

Rosário Oliveira

Graça Cabeçadas

Phytoplankton community structure defined by key environmental variables in Tagus estuary, Portugal

Keywords: Phytoplankton structure, nutrients, salinity, spatial scale, productive period, estuaries

In this work we analyze environmental (physical and chemical) and biological (phytoplankton) data obtained along Tagus estuary during three surveys, carried out in productive period (May / June / July) at ebb tide. The main objective of this study was to identify the key environmental factors affecting phytoplankton structure in the estuary. BIOENV analysis revealed that, in study period, temperature, salinity, silicate and total phosphorus were the variables that best explained the phytoplankton spatial pattern in the estuary (Spearman correlation, ρ=0.803). A generalized linear model (GLM) also identified salinity, silicate and phosphate as having a high explanatory power (63%) of phytoplankton abundance. These selected nutrients appear to be consistent with the requirements of the dominant phytoplankton group, Baccilariophyceae. Apparently, phytoplankton community is adapted to fluctuations in light intensity, as suspended particulate matter did not come out as a key factor in shaping phytoplankton structure along Tagus estuary.

Introduction

Phytoplankton composition is considered as a natural bioindicator because of its complex and rapid responses to fluctuations of environmental conditions [START_REF] Livingston | Eutrophication processes in coastal systems: origin and succession of plankton blooms and effects on secondary production in Gulf Coast estuaries[END_REF]. The main environmental factors recognized as controlling community structure of phytoplankton are physical, (mixing of water masses, light, temperature, turbulence and salinity) and chemical (nutrients). Regulation of phytoplankton in estuarine systems is more complex due to the interaction of freshwater inputs and tidal energy [START_REF] Alpine | Trophic interactions and direct physical effects control phytoplankton biomass and production in an estuary[END_REF]. Further, the increasing human population densities and industrial development have contributed to the eutrophication of these ecosystems.

Tagus estuary, the major Portuguese estuarine system, represents an important ecological system, still subjected to considerable urban and industrial pressures.

Regarding phytoplankton community of Tagus estuary, [START_REF] Cabrita | Spatial and temporal variation of physico-chemical conditions and phytoplankton during a dry year in Tagus estuary (Portugal)[END_REF] conducted studies on biomass (chlorophyll a), [START_REF] Cabeçadas | Phytoplankton production in Tagus estuary (Portugal)[END_REF] investigated in lower estuary photosynthetic response of phytoplankton and [START_REF] Oliveira | Fitoplâncton. In Qualidade ambiental dos estuários do Tejo e Sado[END_REF][START_REF] Oliveira | Fitoplâncton. Estuário do Tejo[END_REF] studied algal composition and abundance along the estuary. More recently [START_REF] Gameiro | Variability in chlorophyll and phytoplankton composition in an estuarine system[END_REF] focused on the annual variability of phytoplankton composition and biomass, as well as physical and chemical parameters, in the restricted upper estuary. A screening model developed by [START_REF] Ferreira | Does biodiversity of estuarine phytoplankton depend on hydrology?[END_REF] derived some general features relating phytoplankton species composition to hydrology of a number of estuarine systems including Tagus estuary, giving an indication of the respective potential biodiversity. On the basis of physical, chemical and phytoplankton biomass a recent study, carried out in Tagus estuary, allowed distinct habitats to be identified in winter and summer. In the latter period the increment of productivity was particularly noticeable in mid-estuary region [START_REF] Brogueira | Identification of similar environmental areas in Tagus estuary by using multivariate analysis[END_REF].

Environmental and phytoplankton data presented in this work were obtained during three surveys along the Tagus estuary carried out in the productive period, providing an opportunity to explore potential reasons for differences in community composition.

Specifically, the aim of this study is to evaluate the phytoplankton dynamics and the environmental variability along Tagus estuary and investigate the main environmental factors affecting phytoplankton structure.

Materials and methods

Area description and sampling

Tagus estuary is one of the largest estuaries in the western European coast, covering an area of 320 km 2 . The estuary is meso-tidal and its shallow upper part is occupied by extensive tidal flats and salt marshes (Fig. 1). The main freshwater source is Tagus river, having an annual average flow of approximately 400 m 3 /s , although its discharge varies greatly from summer to winter between approximately 30 m 3 /s in a dry summer and 2000 m 3 /s in a wet winter (SNIRH, 2006). Tagus river is the main nutrient source to the estuary, but Sorraia and Trancão rivers also represent significant nutrient contributors to the system. Although organic loadings to the estuary have been reduced in recent years, partially treated or untreated effluents from greater Lisbon still enter the estuary, particularly from southern shoreline, contributing approximately 30% for total organic nitrogen and 10% for nitrate (IST, Maretec, 2002).

A total number of 19 stations were sampled in July 2001, May 2002and June 2003 along Tagus estuary (Fig. 1), always during ebb tide.

Surface water samples were collected with a Niskin bottle for measurement of dissolved oxygen (DO), nutrients (nitrate+nitrite referred as NO 3 , ammonium (NH 4 ), phosphate (PO 4 ), silicate (Si(OH) 4 ), total nitrogen (TN), total phosphorus (TP), suspended particulate matter (SPM), and phytoplankton identification and abundance determination.

Physical and chemical measurements

Temperature (T) and salinity (S) were measured in situ utilizing an Aanderaa probe (CTD sensor), calibrated for salinity with an Autosal salinometer. Samples for DO determination were collected and analyzed according to Winkler method as described by [START_REF] Aminot | Manuel des analyses chimiques en milieu marin[END_REF]. Water samples for nutrient determinations were immediately filtered through AcetatePlus MSI (MicroSeparation Inc.) filters and placed in a freezer. Samples for TN and TP measurements were preserved with sulfuric acid (4M) and maintained at 4ºC until analysis (maximum 48h).

Nutrient analysis was carried out using a TRAACS autoanalyser following colorimetric techniques outlined by the manufacturer. Estimated precision (10 replicates) are the following: 2.6% for NO 3 , 2.9% for PO 4 , 0.2% for Si(OH) 4 , 2.0% for NH 4 at middle scale concentrations. Accuracy of nutrient measurements was maintained by using CSK Standards (WAKO, Japan). TN and TP were determined, after simultaneous oxidation according to the ISO methodology (1997) based on Koroleff method (1983), also automatically. SPM was determined gravimetrically by filtering water samples through pre-combusted Whatman GF/F filters and drying at 70 ºC.

Phytoplankton

Phytoplankton samples (500 mL) were preserved immediately after collection with acidified Lugol. Identification was performed using an Axioscop Zeiss microscope with phase contrast. Counting was done according to [START_REF] Utermohl | Zur Vervollkommung der quantitativen Phytoplankton -Methodic[END_REF] and [START_REF] Lund | The inverted microscope method of estimating algal numbers and the statistical basis of estimations by counting[END_REF] with an Invertoscop Zeiss IM 25. Species richness (S), expressed as species number per sample [START_REF] Pielou | Ecological diversity[END_REF], was estimated considering all species identified during the study period. Species diversity was calculated according to [START_REF] Shannon | The mathematical theory of communication[END_REF] (H´=p i log p i , p i =N i /N, N i number of cells of species i and N total number of cells/sample) and Evenness (J´) as H´/log S [START_REF] Pielou | Ecological diversity[END_REF], using log 2 in both formulations.

Statistical Analysis

Multivariate Analysis

Multivariate statistical analysis was applied to environmental and phytoplankton data in order to identify spatial trends or discrete groups and correlations between environmental and phytoplankton variables. These analyses were undertaken using PRIMER (v.6) software (Plymouth Routines in Multivariate Ecological Research, Plymouth Marine Laboratory, Plymouth, U.K; [START_REF] Clarke | Primerv6: User Manual/Tutorial[END_REF].

ANOSIM test

The analysis of similarity, ANOSIM [START_REF] Clarke | Statistical design and analysis for a 'biological effects' study[END_REF][START_REF] Clarke | Non-parametric multivariate analyses of changes in community structure[END_REF]) was performed to test statistical differences in environmental and phytoplankton data among sampling months and sampling stations. We used a two-way crossed test, based on Bray-Curtis similarities (phytoplankton abundances, fourth-root transformed) and Euclidean distance matrix (environmental data, log(x+1) transformed) as our study focused on a fixed set of sites, sampled at several occasions.

PCA, MDS and BIOENV analysis

Environmental data (log(X+1) transformed) was handled using correlation-based Principal Component Analysis (PCA), on the basis of standard Euclidean distance between samples to define their dissimilarity.

Multidimensional scaling, MDS, was applied to phytoplankton abundance (4th root transformed). Similarities between species are obtained by Bray-Curtis similarity coefficient and the corresponding rank similarities used to construct MDS configuration. Sampling points are arranged in a continuum such that points emerging together correspond to sites in which species composition is similar, and points which are far apart correspond to sites that are dissimilar. Stress levels of MDS representation less than 0.1 indicate good representation of the data. Also on the basis of Bray-Curtis similarities, the similarity percentages analysis (SIMPER) was applied to phytoplankton species abundance, in order to allow the separation of every two groups of stations according to phytoplankton species.

Further, patterns in community structure identified by MDS analyses were linked to environmental variables (based on Euclidean similarity index) by using the BIOENV method. This procedure allows identifying the environmental variables (individual or combined) that "best match" the patterns of community structure.

Univariate Analysis

Generalized Linear Model (GLM) was also applied to further explore the relationship between the environmental variables and phytoplankton abundance. This model, adequate for data with counts, uses the method of least squares, by linking the binary response to the explanatory covariates through the probability of either outcome, which does vary continuously from 0 to 1. The transformed probability is then modeled with an ordinary polynomial function, linear in the explanatory variables. This univariate analysis was carried out using Brodgar [START_REF] Software | Brodgar Software for Univariate & Multivariate Analysis, and Multivariate Time Series Analysis[END_REF].

Results

Based on ANOSIM analysis no significant differences among sampling months was found, neither for most environmental data (ρ varying from -0.095 to 0.286, p>5%, except for DO and SPM) nor for phytoplankton data. Therefore, data was averaged over the three months, at each sampling site.

Environmental parameters

Spatial variations of environmental parameters along Tagus estuary during productive period are shown in Fig. 2. Salinity mean values decreased progressively from a a maximum of 34.5 at St.1 in the estuary mouth to a minimum of 0.3 at St. 14, the most river-influenced site (Fig. 2a). An opposite trend was observed in relation to water temperature ranging from 17ºC at St. 1 to 24-25 ºC at Sts. 12-14 (Fig. 2b).

The system was well oxygenated, as oxygen saturation levels remained above 90%, except at St. 15, nearby Montijo saltmarsh, where saturation decreased to 87%, corresponding to 6.5 mg/l, the minimum value found in the estuarine surface water (Fig. 2c).

Nutrient concentrations, corresponding mostly to moderate levels, decreased along the estuary from riverine to marine,influenced area, reflecting in general, the main nutrient discharge from Tagus river. NO 3 concentrations varied from 10 µmol/L at St.1 to 50-60 µmol/L at St.11-14 in the upper estuary (Fig. 2d). The same trend was observed for Si(OH) 4 , ranging from 4 µmol/L to 60 µmol/L (Fig. 2e), and for PO 4 , ranging from 1 µmol/L to 3-4 µmol/L (Fig. 2f). TN distribution also reflects the major input from Tagus river (maximum of 102 µmol/L at St. 14), as well as the influence of the extensive salt-marshes in the upper estuary (St. 8, 9 and 11) (Fig. 2g). TP also varied from minimum values of 2-3 µmol/L in the lower estuary (Sts. 1-3) to 6-7 µmol/L in the upper estuary (Sts. 13, 14) (Fig. 2h).

The existence of NH 4 point sources in the estuary is notorious, maximum value (60µmol/L) being found at St. 15 (nearby Montijo saltmarsh) (Fig. 2i). However, high values were also detected close to St. 16 (Seixal saltmarsh) and St. 4 (nearby Trancão river discharge), respectively 18 and 13 µmol/L. SPM varied quite irregularly from 21 to 58 mg/L all over the estuary (Fig. 2j). Besides St. 4 and St. 5 located in the vicinity of Trancão river discharge, which showed the highest levels of SPM, sites at the left margin of the estuary (Sts 15,16,17,18,19) close to extensive saltmarsh zones, exhibited considerably high variability of suspended matter levels.

Concerning nutrient stoichiometry, it can be observed (Fig. 3) that in the lower/middle estuary (Sts. 1-8 and Sts.15-19) N:P and Si:P values were, in general, lower than 16 and Si:N values lower than 1. By contrast, in the upper estuary (Sts. 9-14) N:P values were in general balanced, some Si:P were above 16, while Si:N were close or above 1.

Phytoplankton Community Structure

Composition and abundance

Mean phytoplankton abundance ranged from a minimum value of 78 cells/mL at St. 18 in the lower/middle estuary to a maximum of 1600 cells/mL at St.11 in the upper estuary (Fig. 4a). Bacillariophyceae was the most important algal group accounting for 87% of the total abundance (Table 1). The algal maximum occurring at St. 11 was mostly due to proliferation of Chaetoceros socialis and Melosira moniliformis (Fig. 4b,c). Two other phytoplankton peaks were observed, one at St. 14 due to proliferation of Stephanodiscus hantzschii (Fig. 4d) and the other at St. 8, mainly as consequence of the development of Chaetoceros socialis and Skeletonema costatum (Fig. 4b,e). This last euryhaline species, adapted to a large salinity range, is even developed at the upper estuary, where the community was characterized by freshwater diatoms Stephanodiscus hantzschii, Fragilaria crotonensis, Aulacoseira distans and A. granulata. At lower and middle estuary other important Bacillariophyceae species were recorded, namely Chaetoceros subtilis, Asterionellopsis glacialis, Cylindrotheca closterium, Detonula pumila and Thalassionema nitzschioides.

Chlorophyceae attained 5% of the total algal abundance (Table1). Its maximum was reached at St. 14, the most river-influenced site (salinity 0.30) being represented mainly by freshwater species (Scenedesmus acuminatus, S. armatus, Monoraphidium contortum, Chlamydomonas spp.). Cryptophyceae were important only at Sts. 4 and 6 (lower/middle estuary), accounting for 4% of the total algal abundance through proliferation of Plagioselmis sp. and Hilea fusiformis.

The oceanic Prasinophyceae, representing 2% of total abundance, developed mainly at the estuary mouth (St. 1) (Table 1) and its principal species was Phaeocystis pouchettii.

Euglenophyceae accounted only for 1% of total abundance, being mostly due to the development of Eutreptiella marina in the lower estuary. The contribution of Cyanobacteria and Dinophyceae and Ebriidea was below 1%. The most abundant dinoflagellates were nanoplanktonic Amphidinium species (Sts. 1 and 6) and the most common Cyanobacteria was Merismopedia tenuissima (St. 12).

Diversity

A total of 236 phytoplankton species were identified in Tagus estuary during the study period. However, species richness (S) was moderate and no clear trend was apparent along the estuary (Fig 5). The minimum value (25 species) was detected at Sts. 3, 18 and 19 in the euhaline -mesohaline zone, and the maximum (37 species) was detected at St. 10 in the oligohaline-freshwater zone (Fig. 5), indicating the importance of freshwater species in the estuarine community. Shannon diversity (H´) presented high values, most of them (84%) over 3.0, ranging from 1.7 (St. 11) to 4.6 (Sts. 6 and 17) (Fig. 5). Evenness (J´) was also high, and 79% of values surpassed 0.6 (Fig. 5). Changes in J´ closely mirrored changes in H´, varying from a minimum of 0.4 at St. 11 to a maximum of 0.9 at St. 17. The occurrence of a simultaneous reduction of H´ and J´ at St. 11, where phytoplankton maximum was attained, reflects the strong dominance of Chaetoceros socialis and Melosira moniliformis (Fig. 4b,c).

Statistical Analysis

The application of PCA analysis to environmental data reveals that the first three PCs had eigenvalues greater than one (Table 2), indicating that all of them were significant.

These components accounted for 89% of total variance, and represent a very good description of the environmental structure.

Variables T, NO 3 , PO 4 , Si(OH) 4 , TN, and TP present highest positive loads to PC1 while S loads negatively (Table 3). This component, accounting for 59.7% of total variance (Table 2) is interpreted as the gradual mixing of nutrient-enriched fresh water with nutrient-poor saline water along the estuary (Fig. 6a,b). Position of most of stations along this axis gives support to this interpretation, St 14 (the most riverine) and St. 1 (at the estuary mouth) being more distant from each other.

PC2 accounted for 19% of total variance and variables NH 4 and DO are, respectively, the main positive and negative contributors to this axis (Table 3). This can be interpreted as a representation of "poor water quality" in specific areas as opposed to the well oxygenated main body of the estuary. The isolation of St. 15, and to a minor extent St. 4 and St. 16, connected with higher values of NH 4 and lower values of DO, supports this interpretation (Fig. 6a,b). The remaining environmental variable, SPM, contributes mostly to PC3 which still explains 10% of variation (Table 3). This component is mainly associated with higher turbidity at Sts. 4 and 5, both located nearby Trancão river discharge.

The MDS ordination of stations generated by phytoplankton abundance is illustrated in Figure 7. Stress value associated with this 2-dimensional plot is 0.11, revealing that this representation of stations is still sound. However, the 3-dimensional MDS (not shown) presents a lower stress (0.07) indicating that samples are likely to fit more easily into three dimensions. The successive position of the stations reveals algal communities differences along the estuary. St 1 and St 14 are the most distinct sites as they differ both in abundance and species composition (respectively marine and freshwater taxa)

According to cluster analysis (not shown) stations located in lower/middle estuary (St.

1-4 and 15-19) constitute a group (A) characterized by the presence of marine and estuarine species (Thalassionema nitzschioides, Guinardia delicatula, Asterionellopsis glacialis, Nitzschia longissima). Sts. 5-10 and 12, mostly located in the middle estuary, formed another distinct group (B) with a mixed dominance of marine, estuarine and some freshwater species (Melosira moniliformis, Chaetoceros subtilis, Thalassionema nitzschioides, Chaetoceros socialis, Stephanodiscus hantzschii). Group C including Sts 11, 13 and 14 individualize the upper estuary due to higher abundance and prevalence of freshwater phytoplankton species (Stephanodiscus hantzschii, Aulacoseira distans, Cyclotella meneghiniana, Scenedesmus opoliensis). Consequently, Groups A and C present the highest dissimilarity (81%, SIMPER analysis) while the average dissimilarity between groups B-C and A-B attained respectively 65% and 62%.

BIOENV analysis reveals that environmental parameters have a strong correlation with phytoplankton community along Tagus estuary (Table 4). The combination of variables that best explained the phytoplankton pattern in the study period were T, S, Si(OH) 4 and TP (ρ =0.803) although correlation with S alone was nearly as large (0.796). Salinity is a dominant factor and is the single variable that is retained in all of the best 10 sets of variables. SPM is still included in the best 10 results, though the respective group of variables have relatively lower correlation with phytoplankton abundance (ρ= 0.776).

The application of GLM analysis revealed the following relationship between phytoplankton abundance and environmental variables: log (Phyto abundance) = 6.98815 -0.07298log(S) + 0.42153log (PO 4 ) -0.02145log(Si(OH) 4 ) This relationship, which includes S, Si(OH) 4 and PO 4 , has a high explanatory power (63%) for phytoplankton abundance.

Discussion

Both environmental and phytoplankton variables showed an upper-lower gradient along Tagus estuary, despite nutrient point sources other than Tagus river being important inputs to specific areas of the system.

The identified phytoplankton assemblages including, respectively, freshwater, estuarine and freshwater, and estuarine and marine species, were dominated by Bacillariophyceae. The dominance of this group has been already reported by other authors, either along the estuary [START_REF] Oliveira | Fitoplâncton. In Qualidade ambiental dos estuários do Tejo e Sado[END_REF]2003) or in the upper [START_REF] Gameiro | Variability in chlorophyll and phytoplankton composition in an estuarine system[END_REF] and lower estuary [START_REF] Cabeçadas | Phytoplankton production in Tagus estuary (Portugal)[END_REF].

The total number of species (236) identified in this study was lower than that referred by [START_REF] Moita | Checklist of phytoplankton species off Portugal: 70 years of studies[END_REF]. However, these authors also included phytoplankton data from winter period, when Tagus river discharge is higher, and many freshwater species enter the estuary. The majority of these allochthonous species do not remain as components of the phytoplankton community all year round. In fact, in March 2001, after a very rainy winter period, a considerable number of freshwater species (115) were identified all along the estuary [START_REF] Oliveira | Fitoplâncton. Estuário do Tejo[END_REF]. By contrast, in July 2001 (data included in this work) only 28 of these freshwater species survived at the upper estuary. On the other hand, the screening model developed by [START_REF] Ferreira | Does biodiversity of estuarine phytoplankton depend on hydrology?[END_REF], using data from 1929 to 1998, predicted that in Tagus estuary, from a total number of 342 species, 120 were combined riverine and oceanic species, and only 222 were real estuarine. This number compares well with the one found in our study.

The strong correlations revealed by BIOENV analysis support the view that, in the study period, the phytoplankton in Tagus estuary, were more likely controlled by abiotic factors than by biotic mechanisms, such as grazing by zooplankton. Despite data on zooplankton not being available for all the study period, [START_REF] Monteiro | Zooplâncton. Estuário do Tejo[END_REF] reported the occurrence in July 2001 of two zooplankton maxima, respectively, one at the oligohaline-freshwater zone coinciding with the phytoplankton maximum, and the other at the estuary mouth, with zooplankton abundance being low in the other estuarine zones.

In fact BIOENV analysis identified salinity, temperature, silicate and phosphorus as variables explaining the phytoplankton spatial pattern, although salinity alone has an overriding role. Results from the GLM analysis agreed quite well with BIOENV by selecting variables salinity, silicate and phosphate as explanatory variables of phytoplankton abundance in Tagus estuary in the study period. The inclusion of silicate and phosphorus emphasizes the nutrients requirements of the dominant algal group (Bacillariophyceae), and may be related to the apparent unbalanced proportions of nutrients in the estuary, revealed by shortage of Si in the lower/middle estuary and of P in the upper estuary (van der [START_REF] Van Der Zee | Seasonal cycling of phosphorus in Southern Bight of the North Sea[END_REF]Justic et al. 2005). Nevertheless, as nutrient levels in the study period were higher than typical half saturation constants for natural phytoplankton populations, growth of most species was unlikely to be nutrient limited [START_REF] Reynolds | Non-determinism to probability, or N:P in the community ecology of phytoplankton[END_REF][START_REF] Estrada | Effects of pulsed nutrient enrichment on enclosed phytoplankton: ecophysiological and successional responses[END_REF].

Regarding the SPM variable, which varied irregularly along the estuary, concentrations were above 10mg/L which represents a threshold value above which primary productivity is usually inhibited (De Master et. al., 1993;[START_REF] Ragueneau | Biogeochemical transformations of inorganic nutrients in the mixing zone between the Danube River and the North-western Black Sea[END_REF].

However, the apparent reduced light conditions appear adequate for Bacillariophyceae growth. In fact, studies carried out by [START_REF] Cabeçadas | Phytoplankton production in Tagus estuary (Portugal)[END_REF] on phytoplankton photosynthetic responses in Tagus estuary concluded that, in the lower estuary phytoplankton seemed partially adapted to high turbid conditions. It is known [START_REF] Smayda | Community assembly in marine phytoplankton: application of recent models to harmful dinoflagellate blooms[END_REF]) that species of this group are characterized by lower-light requirements and adaptation to fluctuations in light intensity induced by turbulent mixing.

Despite nutrient discharges from point sources in specific areas of Tagus estuary, no evidence of poor water quality relationships with changes in phytoplankton structure could be found. In fact, Cyanobacteria and Dinophyceae, which are dominant in high trophic state ecosystems, as well as Cryptophyceae whose predominance during the summer is known to be an indicator of eutrophication [START_REF] Moncheva | Phytoplankton blooms in Black Sea and Mediterranean coastal Ecosystems subjected to anthropogenic eutrophication: similarities and differences[END_REF], were all minor contributors to the algal community. Indeed, they represented only 1-2 % of the phytoplankton total abundance, suggesting that, at present, cultural eutrophication is not a problem in this estuary. In addition, the high values of Shannon-Wiener index and Evenness confirm that most phytoplankton communities are at equilibrium in Tagus estuary.

This work can be valuable for the outlining of water quality monitoring programs, as it allows to take into account a reduced number of key environmental parameters able to characterize the phytoplankton community structure in Tagus estuary Table 1 Dominant algal groups (cells/mL) along Tagus estuary in the productive period.

Fig. 7

 7 Fig. 7

Table 3 .

 3 Variable loads of the environmental PCA

	Algal group								Stations
		1	2	3	4	5	6	7	8	9	10	11	12	13	14	15	16	17	18	19	%
	Bacillariophyceae	332 147 223 224 238 206 432 833 269 672 1562 612 400 934 165 218 116	71	73	87
	Chlorophyceae	13	3	4	7	19	10	24	7	29	5	48	11	59	229	2	6	0.7	0.6	0.8	5
	Cryptophyceae	15	10	8	110 0.4	97	5	3	0.8	28	0.2	8	0.4	0	11	25	1	0.6	8	4
	Cyanobacteria	0	0	0	0	0	0	0	0	3	0	0	22	0	2	3	0	0	0	0	0.3
	Dinophyceae	15	4	1	5	3	16	5	3	0.2	2	0	3	0	0	4	11	0.3	0.8	0.4	0.8
	Ebriidea	0	0.1	0	4	0	4	2	0.1	0.8	1	1	0.2	0	0.1	0	0.1	1	3	1	0.2
	Euglenophyceae	0.6	1	0.7	1	36	34	2	9	4	1	0.6	0	0.1	0	0	1	4	2	0.7	1
	Prasinophyceae	107	4	0	8	0.8	8	1	2	1	0.5	0	0	0	0	2	14	0	0.6	0	2

Table 4 .

 4 Summary of the 10 best results (Spearman correlation, ρ) obtained from BIOENV analysis of combined environmental variables matching phytoplankton abundance

	Number of Variables	ρ ρ ρ ρ	Variables
	4	0.803	T, S, Si, TP
	1	0.796	S
	3	0.794	T, S, Si
	5	0.789	T, S, NO, Si, TP,
	4	0.784	T, S, NO, Si
	3	0.781	S, NO, Si
	2	0.779	S, Si
	2	0.778	S, NO
	5	0.776	T, S, Si, TP, SM
	2	0.776	S, TP

Note: (S in bold as appeared in all combinations)

Fig. 1

Fig. 2

Fig. 3

Fig. 4

Fig. 5

Fig. 6

Acknowledgements

This work was financially supported by EU, under project P.O.MARE "Caracterização Ecológica da zona Costeira" Contract no. 22-05-01-FDR-00015 and project REBECCA Contract no. SSPI-CT-2003-502158. We are grateful to M. Nogueira, A.P. Oliveira, C. Gonçalves, V. Franco and L. Palma for helping in sampling and for assistance in measurements. We would also like to thank the anonymous reviewers for constructive comments and suggestions.


Luigi Piazzi

David Balata
email: dbalata@biologia.unipi.it

The spread of Caulerpa racemosa var. cylindracea in the Mediterranean Sea: an example of how biological invasions can influence beta diversity

Keywords: beta diversity, biological invasion, Caulerpa racemosa var. cylindracea, sublittoral, Mediterranean Sea

The present study aims to evaluate if invasion of Caulerpa racemosa var. cylindracea can affect both alpha and beta diversity in Mediterranean coastal system. The following hypotheses were tested: i) measures of alpha diversity differ between invaded and non invaded assemblages ii) this pattern is consistent between different habitats, iii) assemblages from habitats which are naturally very different tend to be more similar when invaded compared to non-invaded situations. Results showed that Mediterranean benthic assemblages invaded by C. racemosa var. cylindracea were characterized by lower values of alpha diversity than non-invaded assemblages; moreover, differences between deep (25 m) and shallow water (5 m) assemblages appeared lower in invaded than in non-invaded areas, with a decrease of beta diversity. This homogenisation of habitats was related to a lack of several characteristic species and to a higher abundance of few opportunistic species in invaded areas, particularly turf-forming algae. Moreover, variability in taxa composition was lower in invaded than in non-invaded assemblages in both habitats.

Introduction

Biological invasions represent an important component of human-induced global changes with serious consequences for ecological, economic and social systems (Vitousek et al., 1996;1997;Occhipinti-Ambrogi and Savini, 2003). In marine systems, biological invasions are particularly severe with usually profound impacts on structure and functioning of ecological systems, by reducing native species richness and altering ecological processes (Ribera and Boudouresque, 1995;Grosholz, 2002;Larned, 2003;Ruesink et al., 2006).

The influences of biological invasions on biodiversity and on the integrity of natural assemblages are widely investigated in aquatic systems (Cohen and Carlton, 1998;Carlton, 1999;Scheibling and Gagnon, 2006). Invaders are normally highly competitive (Rejmánek and Richardson, 1996) and their interaction with native species can cause the extinction or a strong reduction of the latter, leading to a decrease of ecosystem biodiversity (Walker and Kendrick, 1998;Bax et al., 2003). Most of the studies related to biological invasions have used species richness (the number of species in a community) and/or alpha diversity (includes both the number of species and the relative abundance of individuals) as a surrogate of biodiversity (Viejo, 1997;Piazzi and Cinelli, 2003;Casas et al., 2004, Buschbaum et al., 2006). However, biological invasions can influence other aspects of biodiversity. It has been shown that invaded habitats tend to become similar independently of their native characteristics (McKinney, 2004;McKinney and Lockwood, 1999). Consequently, the invasion process causes a decrease of between-habitats diversity (beta diversity) (Gray, 1997).

This aspect of biodiversity is considered key to understanding the functioning of ecosystems and to the conservation and sustainable use of biodiversity (Legendre et al., 2005), but it has not been much a focus enough in relation to biological invasions in marine systems.

In the Mediterranean Sea, invasions of the green algae of the genus Caulerpa represent an important threat to the biodiversity of benthic coastal systems (Boudouresque and Verlaque 2002). In particular, Caulerpa racemosa (Forsskål) C. Agardh var. cylindracea (Sonder) Verlaque, Huisman et Boudouresque (Verlaque et al., 2003) (C. racemosa from here after) has rapidly and successfully invaded wide areas of this basin, forming permanent populations and strongly interfering with native species (Piazzi et al., 2001;2005b;Ruitton et al., 2005a).

C. racemosa is able to overgrow other macroalgal species, reaching virtually 100% cover in invaded areas (Piazzi et al., 2001;Balata et al., 2004). This spatial dominance can modify the structure and composition of invaded assemblages and reduce differences between habitats, leading to habitat homogenisation.

The aim of the present study was to evaluate if invasion of C. racemosa can affect both alpha and beta diversity. In order to verify these two models, the following hypotheses were tested: i) measures of alpha diversity differ between invaded and non invaded assemblages ii) this pattern is consistent between different habitats, iii) assemblages from habitats which are naturally very different tend to be more similar when invaded compared to non-invaded situations.

Material and methods

The study was carried out on the rocky coast of Tuscany, in the north-western Mediterranean Sea. In this geographic area Caulerpa racemosa was introduced in 1993 and now the alga colonizes wide zones along the continental and insular coasts from 0 to more than 40 meters in depth (Piazzi et al., 2005b).

We selected two habitats at different depths: 5 m and 25 m. For each of the two habitats, 4 locations about 1000 m 2 wide and with similar exposure to light, currents and wave energies, were chosen a few km apart. For each habitat, 2 locations were invaded by C. racemosa for at least two years, while the other 2 were non-invaded. Invaded and non-invaded locations were interspersed. At each location, two sites of about 100 m 2 wide and tens of m apart were chosen on horizontal substrata. In each site, three replicated quadrats of 400 cm² were collected in November 2005 by scraping the substratum with a hammer and a chisel and preserved in formalin. This sampling date was chosen because it is during the peak in biomass production of C. racemosa, before the resting winter season (Capiomont et al., 2005;Ruitton et al., 2005b). Algae and sessile invertebrates were identified in the laboratory and their According to Whittaker (1972), beta diversity can be measured as the variability in species composition among sampling units for a given area at a given spatial scale. Beta diversity was evaluated through a modified version of the Gower's measure of dissimilarity, as described in Anderson et al. (2006). We chose the Gower's measure of dissimilarity because it explicitly weights an order-of-magnitude change in abundance the same as a change in species composition (Anderson et al., 2006). The Gower's measure of dissimilarity was obtained after transforming the data to x'=log 2 (x)+1 for x>0 and x'=0 for x=0, and excluding double zeros.

Non parametric multivariate analysis of variance (PERMANOVA) was used to test the hypothesis that assemblages differed between invaded and non-invaded areas and between habitats and to detect eventual interactions between these factors (Anderson, 2001). Factors in the analysis were the same as in univariate analysis.

Differences in multivariate dispersion between invaded and not-invaded assemblages were tested by PERMDISP analysis on the original Gower's matrix being changes in compositional variability among groups directly interpretable as a change in beta-diversity among groups (Anderson, 2006).

A 2-dimensional nMDS (non metric multidimensional scaling) based on the original Gower's matrix was used for a graphical representation of the data (Clarke and Gorley, 2001).

The program IndVal (Indicator Value) was used to determine which species were responsible for the differences between Conditions in each Habitat (Dufrene and Legendre, 1997).

Results

A total of 107 taxa were identified, 13 Chlorophyta, 12 Ochrophyta, 72 Rhodophyta, 3 Porifera, 6 Bryozoa, 1 Annelida (Table 1 with, nomenclature authority). The total number of taxa was 62 at 5 m and 82 at 25 m in non-invaded areas and 41 at 5 m and 59 at 25 m in invaded areas. Similarly, the average number of taxa per sample was significantly higher in non invaded than in invaded assemblages and at 25 m compared to 5 m (Fig. 1a). The number of taxa also differed between locations (Table 2). ANOVA detected a significant interaction between Condition and Habitat for evenness (Table 2). SNK test showed that non-invaded assemblages had significantly higher evenness values than invaded assemblages at 5 m but not at 25 m (Fig. 1b). nMDS showed that, when invaded by C. racemosa, assemblages at 5 and 25 m were more similar than when non invaded (Fig. 2), which resulted in a significant interaction between Condition and Habitat (Table 3). The structure of algal assemblages also differed among locations, but samples were more dispersed in non-invaded than in invaded areas in both habitats. PERMDISP showed significantly (F = 14.15, P = 0.001) lower average distance from centroids in invaded (19.2 ±0.8) than in non-invaded areas (24.1 ± 0.9).

The IndVal analysis showed that the species that mostly contributed to separate invaded locations from non-invaded ones, were Womersleyella setacea at 25 m, Pterosiphonia pennata at 5 m and Cladophora prolifera and Halopithis incurva at both depths (Fig. 3).

Other taxa that contributed to differences between non-invaded and invaded assemblages included erect algae (Halimeda tuna, Flabellia petiolata, Corallina elongata), encrusting Bryozoans ( algae ( and Encrusting Corallinaceae) and several turf species, in particular the epilitic Chlorophyta Pseudochlorodesmis furcellata (Table 4).

Discussion

Both alpha and beta diversity were lower in areas invaded by Caulerpa racemosa than in noninvaded areas.

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Furthermore, assemblages at 5 and 25 m habitats were more similar in invaded than non-invaded areas, and variability in taxa composition was lower in invaded than noninvaded areas at both depths.

A reduction of alpha diversity in assemblages invaded by C. racemosa had been already

reported (Piazzi et al., 2001;Balata et al., 2004) and differences between invaded and noninvaded assemblages described in this study are similar to those described previously. These Simmons, 2004). Both increased or decreased species richness could result in a higher homogenisation of habitats. Thus, impact evaluations of biological invasions should consider not only the total number of species present in an area, but also the species identity and their ecological function.

Acknowledgements

We wish to thank M. Incera, I. Bertocci and two anonymous referees for their helpful comments that improved the quality of the manuscript.

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  covers determined as percentage of sampling surface covered by vertical projection of each organism. Alpha diversity was evaluated as the number of taxa per sample and per habitat (point richness and sample richness respectively, sensu Gray 2000) and as evenness (expressed as the Pielou's index). Values of point richness and evenness were analysed by 4-way ANOVAs. with Condition (invaded vs. non-invaded) as fixed factor, Habitat (5 m vs. 25 m) as orthogonal and fixed factor, Location (2 levels) as random factor nested in the interaction Condition x Habitat and Site (2 levels) as random factor nested in Location. Homogeneity of variances was checked by Cochran's C-test. Student Newman Keuls (SNK) test was used for a posteriori multiple comparison of means (Underwood, 1997).

  effects can be related to the rapid stoloniferous growth of the alga(Ruitton et al., 2005b) that can inhibit the growth of other sessile species. Inhibition can also be enhanced through sediment trapping(Piazzi et al., 2005a; Piazzi et al., 2007).%Beta diversity may be caused by random variation in community composition and/or by gradients of environmental factors(Legendre et al., 2005). The latter case applies to the study system, where sAlthough correlative studies such as this cannot identify a cause-effect relationship, the results suggest that Mediterranean habitats invaded by C. racemosa tended to lose their natural patterns of variability, leading to a reduction of beta diversity. The loss of spatial variability in different habitats or regions contributes to what is known as "biotic homogenisation" (McKinney and Lockwood, 1999). Biotic homogenisation is considered one of the most prominent forms of biotic impoverishment worldwide and an important aspect in the ecology of invasions (Smart et al., 2006). A loss of alpha and beta diversity, involving the replacement of ecological specialists by widespread generalists, can cause not only a taxonomical homogenization but also a functional homogenisation, with consequent changes in overall community function and a reduction in ecosystem resilience (Olden et al., 2004). Also, homogenisation can be enhanced if the lost species are not randomly distributed among taxonomic and ecological categories (McKinney and Lockwood, 1999). In the present study, sessile animals were less abundant in invaded than in non-invaded assemblages and Bryozoans were completely absent. Although this relation must be further explored by experimental studies, it represents an important issue related to C. racemosa invasion. In fact, the lack of most sessile animals in invaded assemblages can lead to serious functional alterations of invaded ecosystems. Taxonomic and functional homogenisation can increasing the vulnerability to large-scale events, such as environmental alterations or further biological invasions )Although beta diversity has been less considered than species richness in marine ecological studies, it represents a fundamental aspect of the ecosystems and a good ecological indicator. Biological invasions may cause contrasting effects on assemblages that are difficult to understand, as richness can increase or decrease as a consequence of invasions (Sanchez et al., 2005; * ! + + ,). Increased species richness, however, could be a consequence of ecosystem modification caused by invaders, that promote the colonization of opportunistic species or species characteristic of other habitats (Crooks, 2002; Britton-

  .T.(1999). Scale and ecological consequences of biological invasions in the world's oceans. In O.T. Sandlund, P.J. Schei, A. Viken (Eds.), Invasive species and biodiversity management(pp 195-212). Dordrecht: Kluwer Academic Publisher. ., and Legendre, P.(1997). Species assemblages and indicator species: the need for a flexible asymmetrical approach. Ecological Monographs, 7, 345-366. Gorgula, S.K. and Connell, S.D. (2004). Expansive covers of turf-forming algae on humandominated coast: the relative effects of increasing nutrient and sediment loads. MarineBiology, 145, 613-619. Gorostiaga, J.M., Santolaria, A., Sacilla, A., and Diez, I. (1998). Sublittoral benthic vegetation of the Eastern Basque Coast (N. Spain): structure and environmental factors.

Table 1 .

 1 List of taxa. N-I = non-invaded assemblages; I = invaded assemblages. Habitats at water depths of 5 m and 25 m.

	Taxa

Table 2 .

 2 Results of 4-way ANOVAs on values of species number and evenness. Significant values are in bold. Condition: non-invaded assemblages (N-I) vs. invaded assemblages (I);

	Habitat: -5 m vs. -25 m.
			number of species		evenness
	Source	df	MS	F	P	MS	F	P
	Condition = C	1 1680.33 34.38 0.023	0.109 24.16 0.001
	Habitat = H	1	616.33 12.61 0.004	0.031 83.35 0.008
	Location (C x H) = L (C x H)	4	48.87 10.76 0.003	0.001	1.06 0.436
	Site (L (C x H))	8	4.54 0.45 0.884	0.001	0.64 0.740
	C x H	1	0.75 0.02 0.907	0.143 109.05 0.001
	Residual	32	10.18			0.002
	Total	47
	Cochran 's C -test			0.153			0.178
	transformation			none			none
	SNK test		C: S.E.= 1.427; H: S.E. = 1.427		C x H: S.

Table 3 .

 3 Results of PERMANOVA on the structure of assemblages. Significant effects are indicated in bold.

	Source	df	MS Pseudo-F P		Denominator terms
	Condition = C	1 55608.00	35.64	0.001	L (C x H)
	Habitat = H	1 25456.00	16.32	0.001	L (C x H)
	Location (C x H) = L (C x H)	4 1560.30	5.34	0.001	S (L (C x H))
	Site (L (C x H)) = S (L (C x H))	8	292.27	0.76	0.871	R
	C x H	1 6888.80	4.42	0.005	L (C x H)
	Residual = R	32	386.01
	Total	47

Table 4 .

 4 Results of the Indicator Species Analysis (IndVal). Habitats at water depths of 5 m and 25 m.

	-5m
	Non-invaded assemblages	Invaded assemblages
	Halimeda tuna	Cladophora prolifera
	Pseudochlorodesmis furcellata	Pterosiphonia pennata
	Aglaozonia melanoidea	Halopithys incurva
	Corallina elongata
	Encrusting Corallinaceae
	Dasya rigidula
	Dictyota dichotoma
	Laurencia obtusa
	Herposiphonia secunda
	Symphyocladia marchantioides
	Acrosorium venulosum
	Ceramium flaccidum
	Valonia utricolaris
	Padina pavonica
	Hypnea musciformis
	Feldmannophycus rayssiae
	Gelidium spathulatum
	Peyssonnelia rubra
	-25m
	Non-invaded assemblages	Invaded assemblages
	Pseudochlorodesmis furcellata	Cladophora prolifera
	Encrusting Bryozoans	Halopithys incurva
	Halimeda tuna	Womersleyella setacea
	Apoglossum ruscifolium
	Chondria dasyphylla
	Aglaothamnion tenuissimum
	Flabellia petiolata
	Wrangelia penicillata
	Peyssonnelia rubra
	Sphacelaria cirrosa
	Falkenbergia rufolanosa
	Champia parvula
	Lomentaria chylocladiella
	Encrusting Corallinaceae
	Rhodophyllis divaricata
	Erect Bryozoans
	Crouania attenuata
	Contarinia squamariae
	Ceramium codii
	Chondria dasyphylla
	Ptilothamnion pluma

Ballesteros, E., Calvo, S., Cinelli, F., Cirik, S., Cossu, A., D'Archino, R., Djellouli, S.A., Javel, F., Lanfranco, E., Mifsud, C., Pala, D., Panayotidis, P., Peirano, A., Pergent, G., Petrocelli, A., Ruitton, S., Žuljevi , A., and Ceccherelli, G. (2005b). Invasion of Caulerpa racemosa var. cylindracea (Caulerpales, Chlorophyta) in the Mediterranean Sea: an assessment of the early stages of spread. Cryptogamie Algologie, 26, 189-202. Piazzi, L., Balata, D., Foresi, L., Cristaudo, C., and Cinelli, F. (2007)