Stian Røberg

Kristian Svein

Bjarne Stormo

Landfald

Persistence and biodegradation of kerosene in high-arctic intertidal sediment

Keywords: hydrocarbons, kerosene, bioremediation, intertidal zone, Arctic, Svalbard

) and 78 days (2003). The bulk of added kerosene, i.e. 94 to 98 percent, was lost from the upper 5 cm layer by putatively abiotic processes within 2 days and a residual fraction in the range 0.6 to 1.2 mg per g dry sediment was stably retained.

Concomitant addition of oleophilic fertilizer led to higher initial retention, as 24% of the kerosene remained after 2 days in the presence of a modified, cold-climate adapted version of the well-known Inipol EAP 22 bioremediation agent. In these enclosures, which showed an increase in hydrocarbon-degrader counts from 6.5x10 3 to 4.1x10 7 per g dry sediment within 8 days , a 17% contribution by biodegradation to subsequent hydrocarbon removal was estimated. Stimulation in hydrocarbon-degrader counts in fertilizer-alone control enclosures was indistinguishable from the stimulation observed with both kerosene and fertilizer present, suggesting that the dynamics in numbers of hydrocarbon-degrading bacteria was primarily impacted by the bioremediation agent.

Introduction

The risk of accidental oil spills in the polar regions increases with the extent of human activities. The potentially most devastating incidents are obviously related to production and transport of crude oil, but the hazards of contamination from fuel oils are also increased in connection with activities such as tourist cruises, fisheries and research.

In the year-round low temperatures of the Arctic and the Antarctica, evaporation of even relatively light hydrocarbon fractions is low [START_REF] Fingas | Modeling evaporation using models that are not boundary-layer regulated[END_REF] , while viscosity is increased compared with temperate and warm regions [START_REF] Jarsjö | On the relation between viscosity and hydraulic conductivity for volatile organic liquid mixtures in soils[END_REF]. Thus, the residence time of e.g. stranded spills must be expected to increase. This effect is further enhanced by the fact that the physicochemical processes of oil removal from the shorelines are largely discontinued for major parts of the year, when the beaches and adjacent waters are frozen and ice-covered.

Hydrocarbon-degrading bacteria have been shown to be ubiquitous members of the indigenous bacterial community in soil and aquatic environments, including permanently cold regions [reviews by [START_REF] Atlas | Microbial degradation of petroleum hydrocarbons: an environmental perspective[END_REF], [START_REF] Leahy | Microbial degradation of hydrocarbons in the environment[END_REF], [START_REF] Margesin | Biological decontamination of oil spills in cold environments[END_REF]]. Generally, these organisms constitute a small subset of the total community in pristine environments, but their relative numbers tend to increase by orders of magnitude in hydrocarbon-exposed areas. This response has been shown to be rapid, also at low temperatures [START_REF] Delille | Influence of daylight on potential biodegradation of diesel and crude oil in Antarctic seawater[END_REF][START_REF] Margesin | Biological decontamination of oil spills in cold environments[END_REF]. In natural soil and aquatic environments, the hydrocarbon-induced bacterial proliferation is generally restricted by too low inherent bioavailability of the macro-nutrients nitrogen and phosphorous. Growth stimulation by N and P supplementation is well documentedfrequently accompanied by increases in hydrocarbon biodegradation rates [START_REF] Atlas | Degradation and mineralization of petroleum in sea water: limitation by nitrogen and phosphorous[END_REF][START_REF] Gibbs | Quantitative studies on marine biodegradation of oil. I. Nutrient limitation at 14 degrees[END_REF][START_REF] Braddock | Enhancement and inhibition of microbial activity in hydrocarbon-contaminated arctic soils: implications for nutrient-amended bioremediation[END_REF][START_REF] Xia | Enhanced biodegradation of diesel oil in seawater supplemented with nutrients[END_REF] .

Studies on the use of fertilizers to stimulate bioremediation in intertidal zone systems include laboratory simulations [START_REF] Swannell | Evaluation of bioremediation agents in beach microc osms[END_REF][START_REF] Santas | Effects of wave action on the bioremediation of crude oil saturated hydrocarbons[END_REF][START_REF] Bachoon | Microbial community dynamics and evaluation of bioremediation strategies in oil-impacted salt marsh sediment microcosms[END_REF][START_REF] Oh | Effects of nutrients on crude oil biodegradation in the upper intertidal zone[END_REF][START_REF] Röling | Robust hydrocarbon degradation and dynamics of bacterial communities during nutrient-enhanced oil spill bioremediation[END_REF]Xu and Obbard, 2003), smallscale and pilot-scale on-site experiments [START_REF] Venosa | Use of hopane as a conservative biomarker for monitoring the bioremediation effectiveness of crude oil contaminating a sandy beach[END_REF][START_REF] Oudot | Weathering rates of oil components in a bioremediation experiment in estuarine sediments[END_REF][START_REF] Swannell | Impact of bioremediation treatments on the biodegradation of buried oil and predominant bacterial populations[END_REF][START_REF] Choi | Evaluation of fertilizer additions to stimulate oil biodegradation in sand seashore mesocosms[END_REF][START_REF] Delille | Effectiveness of bioremediation of crude oil contaminated subantarctic intertidal sediment: the microbial response[END_REF][START_REF] Prince | Bioremediation of stranded oil on an arctic shoreline[END_REF][START_REF] Pelletier | Crude oil bioremediation in sub-Antarctic intertidal sediments: chemistry and toxicity of oiled residues[END_REF] and evaluations of countermeasures after real spill events [START_REF] Bragg | Effectiveness of bioremediation for the Exxon Valdez oil spill[END_REF][START_REF] Swannell | A field demonstration of the efficacy of bioremediation to treat oiled shorelines following the Sea Empress incident[END_REF][START_REF] Maki | Intrinsic biodegradation of heavy oil from Nakhodka and the effect of exogenous fertilization at a coastal area of the Sea of Japan[END_REF]. The intertidal zone poses particular problems of maintaining sufficiently high levels of added macro-nutrients. Specialized slow-release inorganic or oleophilic organic fertilizers, as well as repeated addition of conventional agricultural fertilizers, have been employed as bioremediation strategies. As comprehensively reviewed [START_REF] Swannell | Field evaluations of marine oil spill bioremediation[END_REF][START_REF] Lee | In situ bioremediation of oiled shoreline environments[END_REF][START_REF] Zhu | Literature review on the use of commercial bioremediation agents for cleanup of oil-contaminated estuarine environments[END_REF], nutrient additions seem to accelerate immediate biodegradation rates, but the effect on the overall long-term extent of removal has been more contradictory. The physical nature of the beach substrate, aeration, emulsion formation by wave action [START_REF] Santas | Effects of wave action on the bioremediation of crude oil saturated hydrocarbons[END_REF] and environmental temperature are among the factors that have been reported to interact with nutrient availability.

A dedicated oil bioremediation agent, like the Inipol EAP 22, which was used extensively during e.g. the Exxon Valdez cleanup (Bragg et al., 1994) , combine surfactant properties to counter formation of water-in-oil emulsions, organic N and P sources and a hydrophobic carbon source (oleic acid) to boost bacterial growth [START_REF] Ladousse | Results of 12 years of research in spilled oil bioremediation: Inipol EAP 22[END_REF]. However, the impact of this extra C-source addition on the biodegradation of the oil contamination itself has been questioned. The bacteria, even if carrying hydrocarbon-degrading capacity, complete the consumption of the alternative C sources before they attack the hydrocarbons. This has been demonstrated in pure culture of Marinobacter hydrocarbonoclasticus [START_REF] Rivet | n-Alkane biodegradation by a marine bacterium in the presence of an oleophilic nutriment[END_REF] and with natural assemblages of intertidal bacteria [START_REF] Swannell | Evaluation of bioremediation agents in beach microc osms[END_REF].

The present study addresses the rate and extent of surface layer removal of a relatively light (kerosene type) hydrocarbon contamination contained on a high-arctic intertidal beach location. Particular emphasis is on the swiftness and extent of the growth response by the inherent bacterial community and its capacity to biodegrade the residual contamination with or without amendment with an oleophilic fertilizer / bioremediation agent.

Materials and methods

Field site and sediment characteristics

The field trials were carried out in the periods July-September of 2002 and 2003 at Kapp Wijk in the northern arms of the Isfjorden fjord system on Svalbard (N 78°36.5', E 015°10'). A flat moraine peninsula, with an inlet well protected from high seas, is situated approximately 45 km from the nearest settlement (Longyearbyen) and has generally experienced little impact by human activities, except for occasional boat traffic in the surrounding waters. The grain size distribution in the upper 5 cm of the intertidal beach sediment was determined by serial dry sieving and classification acc ording to the Udden-Wentworth scheme (Table 1). Sediment density was 1.77 g cm -3 and water content at low tide was in the range 9.7-17.8 %, depending on the interval between water retraction and sampling.

Experimental setup

The experimental periods lasted from 9 July to 26 September in 2002 (79 days) and from 7 July to 23 September in 2003 (78 days). Twelve 30 cm (i.d.) x 100 cm PVC tubes were placed vertically in the intertidal zone and covered and lidded with reflective Al foil to prevent from solar heating. They were dug 40 cm into the sediment and refilled to the same level as the surroundings. The tidal water percolated vertically through the enclosure sediment with jus t a few minutes delay relative to the surrounding sea surface oscillations. The tubes´ positioning relative to the sea level was such that the orifice was not overflowed at maximum spring tides, while the sediment surface was still submerged at neap tides (Fig. 1). The enclosures were separated by 2 m and organized into 3 groups of 4 treatments each. After being allowed to settle for 3 days, triplicate enclosures were subjected to the following treatments: (1) no additions (control); (2) addition of 400 g (corresponding to 570 mg cm -2 ) of a 190-250°C bp kerosene (<0.3% aromatics) supplemented with heavier alkanes [3.5% hexadecane (2003 only), 1.7% heptadecane, 2.2% octadecane 1.1% nonadecane, 1.1% eicosane and 1.7% pristane (2,6,10,14tetramethylpentadecane) by weight]. All hydrocarbons were purum grade Fluka products;

(3) Addition of 50 ml of the oleophilic fertilizer/bioremediation agents Inipol EAP 22 (2002) or Inipol + (2003), produced by CECA SA (France); (4) concomitant addition of 400 g of the kerosene mixture and 50 ml Inipol (10 vol% of hydrocarbons). In the 2003 experiment, addition of 50 ml Inipol + was repeated after 43 days. To assure an even distribution in the enclosure surfaces, the additions were done at ebb tide, but while the sediments were still submerged. The upper 5 cm layer was assumed to maintain a stable oxic atmosphere due its coarseness combined with the twice-a-day replenishments by fresh air and seawater through the tidal oscillations.

Sampling

Temperature was measured in the surface sediment inside and immediately outside three randomly chosen enclosures at low tide every sampling date. Duplicate core sediment samples were collected from each enclosure by pressing 50 ml sterile plastic tubes 5 cm into the surface layer. The samples were subsequently transported on ice to a laboratory facility within 4 hours. The material for microbiological analyses was processed immediately, while the remainder was frozen at -20°C.

Hydrocarbon analyses

Two grams of thawed sediment were suspended in 5 ml dichloromethane with 500 µg 1hexadecene added as an internal standard. The suspension was vortexed twice for 15 s with a 5 min interval. After 30 min rest, 3 ml of the clear supernatant were transferred to glass screw-cap tubes and stored at 5 o C. Samples were injected splitlessly by an A200S autosampler (Fisons Inst ruments, Altrincham, UK) onto a CP-Sil-8 CB column (25 x 0.25 mm i.d., Chrompack International, The Netherlands) installed in a Fisons 8060 Gas Chromotograph connected to a Fisons VG Platform II mass spectrometer. The column temperature was maintained at 50°C for 3 min, then increased by 8°C min -1 to 130°C and 1°C min -1 to 140°C. Thereafter, the column temperature was ramped to 280°C (15°C min -1 ) and held there for 5 min. The injector temperature was 230°C, the detector interface temperature 250°C, the ion source temperature 180°C, and the ionizing voltage 70 eV. Qualitative analyses were performed by scanning from 30 to 400 atomic mass units at a scan rate of 0.9 s/scan. Alternatively, identical GC temperature conditions were programmed on an Agilent 6890N gas chromatograph (Agilent Technologies, Palo Alto, USA ) equipped with a similar column (VF-5ms, 30 x 0.25 mm i.d., Varian, Palo Alto, USA) and a flame ionization detector. Individual alkanes were quantitated from MS data by the areas from the total ion current or from flame ionization detector areas.

Microbiological methods

Cell suspensions in phosphate-buffered 2% NaCl (PBS) were obtained according to Anderson et al. (2001) and 10-fold dilution series in PBS were used for viable cell enumerations. Numbers of viable hydrocarbon degraders (DC) were estimated by a modified most probable number (MPN) procedure [START_REF] Brown | Sheen Screen, a miniaturized most-probable-number method for enumeration of oil-degrading microorganisms[END_REF].

Aliquots of 20 µl were added into 7 separate dilution series in 96-microwell plates (NUNC A/S , Denmark). The wells were pre-filled with 180 µl 1.5% NaCl-supplemented Bushnell-Haas mineral medium (Difco) and 5 µl filter-sterilised kerosene mixture. After incubation in the dark at 6 o C for 3 months [START_REF] Delille | Field observations on the variability of crude oil impact on indigenous hydrocarbon-degrading bacteria from sub-Antarctic intertidal sediments[END_REF], each well was inspected for dispersion of the kerosene ("Sheen Screen" ) and subsquently added 50 µl filter-sterilized 3 mg ml -1 solution of the respiration indicator INT (2-(4-iodophenyl)-3(4nitrophenyl)-5-phenyltetrazolium chloride; Sigma-Aldrich) [START_REF] Haines | Measurement of hydrocarbon-degrading microbial populations by a 96-well pla te most-probable -number procedure[END_REF]. After over -night incubation in the dark at 6 o C, red and pink wells were counte d as positive for growth. A maximum-likelihood estimation of microbial numbers based on 7-fold dilution series was developed for the Microsoft Excel ® for Windows spreadsheet program.

Additionally, DC were estimated from colony counts of the appropriate dilutions on Petri dishes with solidified Bushnell-Haas medium after incubation at 6°C in a saturated kerosene atmosphere [START_REF] Dandie | Comparison of a range of green fluorescent protein-tagging vectors for monitoring a microbial inoculant in soil[END_REF].

For heterotrophic-count estimates (HC), dilution series were plated on Petri dishes with a half -strength Marine broth/peptone medium (10 g L -1 Marine broth 2216 (Difco) , 5 g L -1 Bacto-peptone (Difco) 15 g L -1 NaCl, 15 g L -1 agar (Difco). Plates were incubated for 3 weeks at 6 o C. Suspended cells for epifluorescent microscopy total counts (TC) were subjected to short-term fixation in 4% paraformalde hyde with subsequent freeze storage in 50% ethanol according to the procedure of [START_REF] Pernthaler | Fluorescence in situ hybridization (FISH) with rRNA-targeted oligonucleotide probes[END_REF]. The fixated cells were diluted 100-fold in ultrapure water, filtered through Whatman Anodisc 25 filters (0.2 µm pore size) and stained with SYBR Gold ® (Sigma-Aldrich, USA) according to [START_REF] Noble | Use of SYBR Green I for rapid epifluorescence counts of marine viruses and bacteria[END_REF], except that stained filters were dried overnight at 40°C prior to mounting. A total of 500-1000 bacter ia-sized cells per filter were counted in a fluorescent microscope (Leitz DMRBE).

Nitrogen, phosphorous and carbon content

Sediment samples (25 g) were dried for 24 h at 40 o C, passed t hrough a 2 mm mesh sieve and milled. Ten mg material was wrapped in tin cups and analysed for total C and N using a Perkin Elmer 2400 CHN Element Analyzer (carrier gases Ar, O 2 and He).

Samples were heated to 640 o C and 926 o C. Inorganic and organic phosphate were extracted [START_REF] Aspila | Semiautomated method for determination of inorganic, organic and total phosphate in sediments[END_REF] and quantified by the colorimetric method of [START_REF] Murphy | A modified single solution method for determination of phosphate in natural waters[END_REF].

Results

Temperature and nutrients

The surface sediment temperatures within the enclosures, as recorded at the sampling dates (Fig. 2), showed the same basic developments both years, i.e. fluctuations in the range 6 to 13°C in the period from start to mid-August, thereafter decreasing temperatures until freezing conditions emerged towards the end of the experiments. The reflective Al foil cover kept the insides of the enclosures slightly on the cold side of the naturally exposed beach throughout the summer periods. The differences were in the range 0.5 to 1 degree on cloudy days and 2 to 3 degrees in sunshine (data not shown).

The 7.4 weight% urea -N in the Inipol bioremediation agent [START_REF] Swannell | Field evaluations of marine oil spill bioremediation[END_REF] added 5.2 mg cm -2 of nitrogen to the fertilized enclosures. This quantity c losely equalled the cumulative amount of total inherent N per cm 2 in the upper 5 cm of the sediment, as determined from the un-fertilized enclosures of the 2003 experiment (0.60 ± 0.07 mg per g dry sediment). However, no trend towards increased N levels in fertilized enclosures was detectable at first samplings 2-5 days after fertilizer addition or later throughout the experimental period. The added 0.5 mg organic P cm -2 (0.7% by weight as tri(laureth-4)-phosphate in the Inipol agent; [START_REF] Swannell | Field evaluations of marine oil spill bioremediation[END_REF] was a negligible contribution to total P of this exceptionally phosphorite-rich [START_REF] Krajewski | Phosphorite -bearing sequence of the Wilhemöya formation in Van Keulenfjorden[END_REF]) sediment (23 ± 3 mg per g dry sediment). It was equivalent to about one-third of the inherent organic P of the upper 5 cm layer (average 0. 16 mg per g dry sediment). Consequently, no significant differences in total or organic phosphate content were observed between treatments.

Hydrocarbons reduction

The bulk of added hydrocarbons were lost from the surface layer within the first few days. This was most evident in enclosures without accompanying fertilizer. In the 2002 experiment, about 98% of the initial kerosene load of 570 mg per cm 2 was lost from the upper 5 cm of sediment after 2 days, while the corresponding figure for 2003 was 94%.

Thereafter, the hydrocarbon reduction levelled out and residual concentrations stabilized at about 0.6 mg per g dry sediment (2002) and 1.2 mg per g (2003) (Fig. 3). With concomitant addition of the Inipol bioremediation agent, a higher fraction of the kerosene was retained in the upper layer. This was particularly evident in the 2003 experiment, when 24% of the initial load (i.e. 15.0 mg per g dry sediment) was retained at day 2. In 2002 the figure was 4.1%. The subsequent removal rates in the fertilized enclosures were, however, higher, resulting in a merging of the retention curves after about a month (2002) or at the time of the termination of the experiment (2003).

As the present kerosene product, on the whole, contained no biodegradationrecalcitrant components, equal quantities of pristane and heptadecane were included to monitor their ratio change as indicator of biodegradation. None of the kerosenesupplemented enclosures showed reduced ratio in 2002, nor did the un-fertilized enclosures in 2003 (Fig. 4). Rather, the straight-chain alkane tended to be retained in the sediment to a higher degree than its highly branched counterpart. This development seemed caused by physiochemical differences, as comparison of the fate of n-alkanes of various chain lengths confirmed a strict dependence on molecular size (Fig. 5). The C 17 and heavier n-alkanes disappeared more slowly than pristane, while the C 16 and shorter molecules showed the opposite pattern. Thus the division point between molecules showing higher and lower retention than pristane did not coincide with heptadecane, but rather was situated in-between C 16 and C 17 .

For Inipol-supplemented kerosene enclosures in the 2003 experiment, the n-C 17 / pristane ratio decreased throughout the experimental period. An abrupt drop by 20% within the first 5 days (which was also observed in the 2002 experiment) was followed by a steady decline to a final ratio of 0.51 (Fig. 4). By disregarding the immediate reduction, which was plausibly abiotic by nature, a biodegradative contribution to overall n-C 17 loss from the upper layer could be estimated at 17% in the period of day 5 to the termination of the experiment at day 78. This is a minimum estimate, as some accompanying microbial consumption of the pristane may be anticipated.

Microbial enumerations

The confinement of beach sediment in experimental enclosures seeme d, by itself, to have little influence on cell numbers. Total epifluorescent counts (TC), which were in the ranges 3.4 to 6.6 x 10 7 per g dry sediment in 2002 and 1.2 to 2.8 x 10 7 per g in 2003, as well as viable heterotrophic (HC) and hydrocarbon-degrader counts (DC), all showed pronounced stability in the control tubes throughout the experimental periods (Fig. 6).

Most-probable -number (MPN) estimates and colony-forming units on solid alkanesmineral salts media (data not shown) both gave estimates of DC in the range 0.01-0.03% of total cells, while the HC figures, on the whole, were 10-fold higher than DC. Based on colony morphologies and microscopic cell-size estimates , this hydrocarbon-degrading viable flora was exclusively of bacterial origin.

Addition of the kerosene mixture led to a marked increase in DC in the surface 5 cm layer within the first 2 weeks. For the rest of the experimental periods, the alkane degraders constituted a major part of viable cells, however still a minor fraction of the total counts. At most samplings, the HC and DC figures were two orders of magnitude below TC. The TC development further manifested the rather insignificant growthstimulatory effect of hydrocarbons alone, as no increase relative to the controls was traceable.

In the first-year experiment, the lack of growth stimulation was evident even when 10 vol% of the Inipol EAP 22 agent was added concomitantly with the kerosene.

However in 2003, the fertilizer-supplemented tubes showed a more notable development.

Both the HC and DC figures exceeded 10 7 per g dry sediment within a week, making viable cells a substantial fraction of the total microbial load. The growth-stimulatory effect was also reflected in a 3-fold increase in total counts. On the other hand, the stimulation was distinctly transient. By the next viable -count sampling 5 weeks later, the DC was down by two orders of magnitude. Following a second fertilizer addition, the high load of viable cells recurred. A six -fold increase in TC, peaking at 1.3 x 10 8 per g dry sediment within 6 days, seemed exclusively due to the growth of readily cultivable variants -including a substantial fraction with capacity for alkanes catabolism.

The boost of alkane degraders in the presence of the Inipol agent in 2003 was evidently effected by components of this product. When added without accompanying hydrocarbons, the oleophilic fertilizer caused developments in microbial numbersalkane degraders included -which basically were indistinguishable from the ones observed in the kerosene + fertilizer enclosures.

Discussion

The 94 to 98% loss of kerosene from the upper 5 cm layer within 2 days in unfertilized enclosures was evidently a result of physicochemical processes, i.e. a combination of leaching deeper into the sediment and evaporation. The laboratory studies of [START_REF] Jarsjö | Retention and volatilization of kerosene: laboratory experiments on glacial and postglacial soils[END_REF] indicate an inherent retention capacity of less than 10 mg per g dry sediment in this rather coarse beach sediment and a shift towards heavier alkanes in the residual fraction (data not shown) indicated a substantial loss also from evaporation. The residual fraction, seemingly recalcitrant to both evaporation and further leaching, was in the range 0.6-1.2 mg per g dry sediment.

Concomitant addition of the Inipol bioremediation agent enhanced retention of the kerosene in the surface layer. The higher hydrocarbon concentration persisted throughout the experimental period in the 2003 experiment, while the residual levels in fertilizer-amended and un-amended enclosures merged after about one month in 2002.

This difference suggested an impact of the replacement of the well-established Inipol EAP 22 product in 2002 with Inipol + in 2003. Basically, the relative content of the cosurfactant 2-butoxyethanol had been increased relative to the other ingredients to obtain a formulation that worked more efficiently than EAP 22 at low environmental temperatures (F. Périé, personal communication).

The study confirmed the presence of a significant indigenous population of hydrocarbon degraders, also in remote polar beach sediments supposedly little affected by human activities [START_REF] Swannell | Evaluation of bioremediation agents in beach microc osms[END_REF][START_REF] Delille | Field observations on the variability of crude oil impact on indigenous hydrocarbon-degrading bacteria from sub-Antarctic intertidal sediments[END_REF]. Addition of the kerosene mixture caused rapid and lasting 1-2 orders of magnitude increases in DC, making hydrocarbon degraders a predominant element of the cultivable bacteria . A highly similar response has previously been observed in sub-Antarctic intertidal sediments [START_REF] Delille | Field observations on the variability of crude oil impact on indigenous hydrocarbon-degrading bacteria from sub-Antarctic intertidal sediments[END_REF]. The attained DC levels of 10 5 -10 6 per g dry sediment were in the same range as reported for unfertilized plots of comparable field experiments [START_REF] Delille | Field observations on the variability of crude oil impact on indigenous hydrocarbon-degrading bacteria from sub-Antarctic intertidal sediments[END_REF][START_REF] Delille | Effectiveness of bioremediation of crude oil contaminated subantarctic intertidal sediment: the microbial response[END_REF] , however generally lower than what has been observe d in laboratory and microcosm simulations of intertidal-zone biodegradation [START_REF] Swannell | Evaluation of bioremediation agents in beach microc osms[END_REF][START_REF] Santas | Effects of wave action on the bioremediation of crude oil saturated hydrocarbons[END_REF][START_REF] Oh | Effects of nutrients on crude oil biodegradation in the upper intertidal zone[END_REF]. Even after growth stimulation, hydrocarbon degraders constituted just 1-2% of total epifluorescent counts in the present study. There is ample evidence that substantial fractions of fluorescently labelled bacteria from natural marine environments are inactive or even dead [START_REF] Lebaron | Does the high nucleic acid content of individual bacterial cells allow us to discriminate between active cells and inactive cells in aquatic systems?[END_REF][START_REF] Luna | Large fraction of dead and inactive bacteria in coastal marine sediments: Comparison of protocols for determination and ecological significance[END_REF]. Consequently, the contribution by the hydrocarbon degraders to the actively metabolizing subset of microbes may have been substantially higher. On the other hand, the observation that total counts were basically unaffected by kerosene -alone addition both years, pointed to a modest impact on the overall biomass. Thus, the microbial growth seemed severely restricted by the availability of inorganic nutrients. Additionally, loose attachment of the hydrocarbondegraders to the substrate may have resulted in repeated reductions in numbers by flushing out of the surface layer during tidal cycles. The low impact of the hydrocarbondegrading flora was also manifested by the striking stability of the residual kerosene fraction after the initial, supposedly abiotic reduction. In 2002, this stability was complete from two weeks on, while a minor reduction was observed during the last month of the 2003 experimental period. Addition of Inipol EAP 22 did not lead to stimulation in hydrocarbon-degrader counts beyond that of kerosene alone in the 2002 experiment. This contrasts previous studies, where experimental units amended by the product typically have reached 1-2 orders of magnitude higher viable counts than un-fertilized controls [START_REF] Swannell | Evaluation of bioremediation agents in beach microc osms[END_REF][START_REF] Choi | Evaluation of fertilizer additions to stimulate oil biodegradation in sand seashore mesocosms[END_REF][START_REF] Delille | Effectiveness of bioremediation of crude oil contaminated subantarctic intertidal sediment: the microbial response[END_REF]. Seemingly, insufficient quantities of the bioremediation agent resisted drainage out of the upper layer by the tidal flushings to make a difference with respect to microbial growth. Inipol EAP 22's susceptibility to leaching from irrigated sediments has been documented in other studies (Xu and Obbard, 2003). As experimental conditions -temperature regime included -were virtually identical the two years, the distinctly stronger growth stimulation by the Inipol + applied in 2003 seemed ascribable to the above-mentioned adjustment in the recipe of the product. Most likely, more of the easily accessible organic nutrients were retained in the surface layer.

Both years the growth stimulation in enclosures added just the fertilizer, was basically identical with what was observed in enclosures added Inipol and kerosene. The 2003 experiment, in particular, suggested that the numerical dynamics of the hydrocarbon-degrading community was distinctly more impacted by the bioremediation agent than by the hydrocarbons per se. One claimed effect of the Inipol products is to initiate an early start of bacterial growth, implicitly of strains with the capacity to catabolize hydrocarbons. This characteristic has primarily been attributed to the presence of about 25% oleic acid [START_REF] Ladousse | Results of 12 years of research in spilled oil bioremediation: Inipol EAP 22[END_REF]. As mentioned, the benefit of this growth boost for subsequent biodegradation has been questioned [START_REF] Rivet | n-Alkane biodegradation by a marine bacterium in the presence of an oleophilic nutriment[END_REF][START_REF] Swannell | Evaluation of bioremediation agents in beach microc osms[END_REF][START_REF] Pelletier | Crude oil bioremediation in sub-Antarctic intertidal sediments: chemistry and toxicity of oiled residues[END_REF]. From the 2003 results of the present study, even the ability of the Inipol-induced flora to sustain on the hydrocarbons when more readily degradable C sources are depleted seems limited. The rapid alkane-degrader proliferations observed after each Inipol + addition were equally transient both if the product was added as kerosene supplement or if it was added to the enclosures alone.

Decline over time in the quantitative ratios between the isoprenoid hydrocarbons pristane and phytane and the respective n-alkanes heptadecane and octadecane has a long record as indicators of biodgradation in oil [START_REF] Blumer | Oil pollution : persistence and degradation of spilled fuel oil[END_REF] , and the general notion that reductions in these ratios are true reflections of biodegradation still prevails.

However, there have also been repeated warnings that quantitative conclusions, at least should be drawn with caution. Both bacterial isolates and consortia have been shown to catabolize the highly branched isoprenoid chains, although mostly at lower rates than straight-chain alkanes [START_REF] Jobson | Mic robial utilization of crude oil[END_REF][START_REF] Oudot | Weathering rates of oil components in a bioremediation experiment in estuarine sediments[END_REF][START_REF] Garrett | Biodegradation of fuel oil under laboratory and arctic marine conditions[END_REF]Xu et al., 2003).

When the ratios between presumed biodegradable and recalcitrant molecules are used as indicator of biodegradation, the possibility of abiotically induced ratio change is implicitly disregarded. To our knowledge, the validity of this premise has not been thoroughly re-examined in available literature. From the present study, however, it seems questionable with respect to the heptadecane / pristane pair. Both years the ratio dropped by 20% of its original value within the first 5 days in Inipol-amended enclosures. These periods were too short in time and alkane-degrader loads were too modest to make biodegradation a plausible explanation. As corresponding developments were not observed in the kerosene -only enclosures, this immediate reduction seemed caused by interaction between the hydrocarbons and the bioremediation agent. Furthermore, the removal rates of the n-alkanes relative to pristane in the subsequent weeks were sensitive to carbon chain length in a manner that seemed incompatible with microbial degradation and the heptadecane and pristane retentions did not coincide. Rather, the removal of the straight-chain hydrocarbon proceeded somewhat slower.

In the light of the above-mentioned, putatively non-biotic influences on the C 17 / pristane ratio, the steady decline in the 2003 kerosene + Inipol enclosures throughout the experimental period suggested some contribution by biodegradation to the overall hydrocarbon removal. The stronger stimulation in alkane -degrader counts, with peak MPN figures exceeding 10 8 per g dry sediment , pointed in the same direction. The minimum estimate of 17% biodegradative contribution to the overall removal of residual hydrocarbons is, however, a modest figure compared with previous observations in intertidal sediments [START_REF] Swannell | Evaluation of bioremediation agents in beach microc osms[END_REF][START_REF] Wright | Oil bioremediation in salt marsh mesocosms as influenced by N and P fertilization, flooding, and season[END_REF][START_REF] Delille | Effectiveness of bioremediation of crude oil contaminated subantarctic intertidal sediment: the microbial response[END_REF].

Conclusions

The present study addressed the fate of a light fuel oil (kerosene) spill in high-arctic intertidal-zone sediment and the biodegradative response by the inherent microbial community. The major findings were as follows:

(1) In the upper 5 cm layer, a residual fraction in the range of 0.6-1.2 mg per g dry sediment seemed recalcitrant to abiotic removal processes throughout experimental periods of nearly 80 days.

(2) Concomitant addition of an oleophilic fertilizer / bioremediation agent slowed the immediate abiotic removal, leaving substantially higher concentrations of the kerosene in the upper layer for weeks or months.

(3) In fertilizer-amended experimental enclosures, the dynamics in numbers of hydrocarbon-degrading cells seemed primarily impacted by the fertilizer; the dynamics in control enclosures with fertilizer only was indistinguishable from what was observed with both kerosene and fertilizer added.

(4) Use of the ratio between heptadecane and pristane as biodegradation indicator was obscured by evidently physicochemical differences in removal rates.

However, it may be concluded that biodegradation was negligible in un-fertilized experimental enclosures. Fertilization gave mixed results, as the established commercial variant of the employed bioremediation agent caused no observable biodegradation, while a minimum 17% contribution by microorganisms to overall hydrocarbon removal was estimated with a modified, cold-climate adapted variant of the product.

Aknowledgements

  This work has been financed by grants from Total E&P Norway AS and the Norwegian College of Fishery Science at the University of Tromsø. The study was also supported financially by the Roald Amundsen Centre for Arctic Research and the Kellfrid and Helge Jacobsen Foundation of the University of Tromsø. Laboratory facilities during fieldwork were provided by the University Centre in Svalbard. The Inipol bioremediation products were kindly provided by CECA SA, France.

Figure legends Figure 1

 legends1 Figure legends

Figure 2

 2 Figure 2 Sediment surface temperatures within enclosures in the experimental periods 9

Figure 3

 3 Figure 3 Total residual extractable hydrocarbons in the enclosures throughout the

Figure 4

 4 Figure 4 Development in the heptadecane / pristane ratios throughout the experimental

Figure 5

 5 Figure 5 Ratios between individual n-alkanes and pristane in kerosene-only enclosures of the 2003 experiment. The values in the original kerosene mixture were

Figure 6

 6 Figure 6 Microbial enumerations by epifluorescent microscopy counts (?), viable

Figures

  Figures A -D give the 2002 data (9 July to 12 September) and figures E-H the

Table 1 .

 1 Sediment grain size distribution and classification (according to the Wentworth scale) of three individual samples distributed throughout the experimental field.

	Classification	Particle size	% of dry sediment
		(mm)	(S.D)
	Medium gravel	8 -16	6.3 (4.9)
	Fine gravel	4 -8	31.3 (6.0)
	Very fine gravel	2 -4	17.9 (2.6)
	Very coarse sand	1 -2	11.9 (3.0)
	Coarse sand	0.5 -1.0	14.8 (1.1)
	Medium sand	0.25 -0.50	13.7 (4.0)
	Very fine sand	0.063 -0.125	0.3 (0.0)

Fig.1

Fig.2

Fig.3

Fig.4

Fig.5

Fig.6


Ben Kahla

Nakbi Amel
email: nakbi_amel@yahoo.fr

Besbes Amine

Chaieb Kamel

Rouabhia Mahmoud

Bakhrouf Amina

Ben Kahla-Nakbi Amel

Amel Ben Kahla-Nakbi

Survival of Vibrio alginolyticus in Seawater and Retention of Virulence of Its Starved Cells

Keywords: Vibrio alginolyticus, survival, seawater, VBNC, ARDRA, virulence

The aim of this study was to evaluate the survival responses of four strains of Vibrio alginolyticus in seawater under starvation conditions. We used microcosms containing sterilised seawater and incubated at ambient temperature (22-25°C). Vibrio alginolyticus maintained its culturability for at least nine months. Long-term-starved cells showed an absence or a decrease in their enzymatic activities. Resuscitation assays of viable but nonculturable (VBNC) cells were conducted and the recovery of these cells was achieved after the addition of nutrients. Amplified 16S ribosomal DNA (rDNA) restriction analysis (ARDRA) was used to confirm that the same strain of Vibrio alginolyticus persisted in all microcosms during a long period of time. Starved cells maintained their infectivity for gilt head sea bream (Sparus aurata) and sea bass (Dicentrachus labrax) as determined by intraperitoneal challenges.

Introduction

Lack of nutrients in natural ecosystems are often the limiting factor for microbial populations.

Under unfavourable environmental conditions, Vibrio spp., like other bacterial species including Aeromonas, Campylobacter, Salmonella and Shigella, have a survival strategy, entering a viable but not culturable (VBNC) state [START_REF] Biosca | Effect of low temperature on starvation-survival of the eel pathogen Vibrio vulnificus biotype 2[END_REF][START_REF] Jiang | Survival of Vibrio parahaemolyticus at low temperatures under starvation conditions and subsequent resuscitation of viable, nonculturable cells[END_REF]. In this state, the cells remain viable but no longer grow in, or develop colonies on, the non-selective media normally employed for their culture [START_REF] Jiang | Survival of Vibrio parahaemolyticus at low temperatures under starvation conditions and subsequent resuscitation of viable, nonculturable cells[END_REF].

In the last decade, it has been frequently reported that under appropriate conditions, these microorganisms are nevertheless able to recover from this dormant state, becoming metabolically active and fully culturable (Nilsson, Oliver, & Kjelleberg, 1991;[START_REF] Whitesides | Resuscitation of Vibrio vulnificus from the viable but nonculturable state[END_REF]). However, it has been proposed that the VBNC state is an adaptative strategy of microorganisms against stress from which cells may be able to recover once optimal conditions are restored [START_REF] Huq | A microbiological paradox: viable but nonculturable bacteria with special reference to Vibrio cholerae[END_REF][START_REF] Nybroe | Assessment of metabolic activity of single bacterial cells -new developments in microcolony and dehydrogenase assays[END_REF][START_REF] Oliver | The viable but nonculturable state in the human pathogen Vibrio vulnificus[END_REF]. This problem is even more complex for bacterial pathogens, since these microorganisms in VBNC state may remain pathogenic [START_REF] Rahman | Methionine uptake and cytopa thogenicity of viable but nonculturable Shigella dysenteriae type 1[END_REF].

Vibrio alginolyticus is a species with a broad geographical distribution in marine and estuarine waters; especially in bathing areas [START_REF] Barbieri | Occurrence, diversity, and pathogenicity of halophilic Vibrio spp. and non-01 Vibrio cholerae from estuarine waters along the Italian Adriatic coast[END_REF][START_REF] Baffone | Occurrence and expression of virulence-related properties of Vibrio species isolated from widely consumed seafood products[END_REF]. It has been associated with several diseases of marine animals, including fish and shellfish (Balebona, Andreu, Bordas, Zorrilla, Moriñigo, Borrego, through seawater, either as bacterioplankton or on the surface of particles. It is also not known whether this bacterium is endemic to the marine environment and whether it behaves as an opportunistic fish pathogen.

The aim of the present work was to investigate the effect of marine stress on the behaviour of V. alginolyticus incubated in seawater microcosms. In addition, to assess the environmental health hazard posed by this bacterium, we studied the ability of long-term-starved cells to express virulent exoenzymes and infect cultured fish.

Materials and Methods

Bacterial Strains and Growth Conditions

Four strains of V. alginolyticus isolates representative of different epizootic events were used in this study. All were recovered from the internal organs of diseased gilt head sea bream (VD4, VD5) and sea bass (VL12, VL13) in a Tunisian aquaculture centre. Stock cultures were frozen at -80°C with 20% (vol/vol) glycerol. Bacteria were routinely cultured on tryptic soy agar ( TSA-1) or in tryptic soy broth (TSB-1) supplemented with 1% NaCl (Difco) at 22°C. Bacterial strains were subjected to standard morphological, physiological and biochemical plate and tube tests using the procedures described in Bergey's Manual of Systematic Bacteriology (Krieg and Holt, 1984). Gram staining, the oxidase test, morphology, motility, susceptibility to the vibriostatic compound O/129 and growth on thiosulfate citrate bile sucrose agar were the main assays employed to identify the organisms. A commercial miniaturized API 2OE Kit (Bio-Mérieux) was also used.

Survival Assay

500 ml of sterile seawater from the Tunisian coast of Monastir (salinity 4%, pH = 8) until a concentration varying between 10 6 and 10 8 colony-forming units (CFU) per ml was reached.

All microcosms were incubated in a static state at ambient temperature (22-25°C).

Enumeration Techniques

Microcosms were sampled daily during the first week, weekly during the first three months, and then once a month. Time zero (inoculation time) and subsequent samples were taken for plate counts. Plate counts of culturable cells were determined by the drop plate method [START_REF] Hoben | Comparison of the pour, spread, and drop plate methods for enumeration of Rhizobium spp. in inoculants made from presterilized peat[END_REF], using marine agar (MA) and selective medium Thiosulfate-Citrate-Bile Salts-Sucrose (TCBS). The plates were incubated at room temperature, and the number of colonies was counted after 24 and 48 h).

Statistical Analysis

Two microcosms per strain were constructed. Three samples per microcosm were taken for plate counts using marine agar (MA) and selective medium (TCBS). Culturable-cell counts from the experiments were processed for variance analysis using the procedure of the general lineal model with repeated measures available in the SPSS package.

.

Resuscitation of VBNC Cells

When cells were considered to be non-culturable (<1 CFU/ml), reversion to a culturable state was attempted by removing 5 ml of seawater from each microcosm at ambient temperature and placing them in a 10 ml Luria -Bertani (LB) broth. After incubation at 28°C, culturability was assayed in marine agar.

Biochemical Activities and Exoenzyme Expression

Biochemical activities of initial cells of V. alginolyticus and long-term-starved cells incubated in seawater was performed on API 20E (Bio-Mérieux). The presence of several exoenzymes was determined using the API Zym System (Bio-Mérieux) composed of 19 enzymatic substrates. The activities of various other enzymes were determined following inoculation of cultures onto TSA-1 to which the following substrates had been added: 0,2% [wt/vol] strach for amylase, 1% [wt/vol] skim milk for caseinase, 1% [wt/vol] gelatin for gelatinase, 1%

Tween 80 for lipase, 5% [vol/vol]) egg yolk for phospholipase (lecithinase) and 5% [vol/vol])

sheep red blood cells for haemolysin.

ARDRA (Amplified Ribosomal DNA Restriction Analysis)

Amplified 16S ribosomal DNA (rDNA) restriction analysis (ARDRA), using enzyme HaeIII was applied to starved cells of V. alginolyticus (cells incubated in seawater microcosms) and initial cells to compare their profiles.

Samples to be used for ARDRA analysis were obtained by centrifuging 5 ml of the starvation micr ocosm at 16,000 g for 5 min. Cells were washed and resuspended in 10 ml of LB broth.

After 24-hour incubation at 28°C, chromosomal DNA was extracted using a Wizard Genomic Purification Kit (Promega). The amplified 16S rRNA gene was obtained from each isola te and reference strain by PCR with the universal primers F (5' AAT CAC TAC CGG AAA CAG CG 3') and R (5' TGC ACT CAA TGA GGG AAT CA 3'), targeted to universally conserved regions and permit the amplification of an approximate 1,500-bp fragment. PCR amplification was performed by 2 µl of bacterial DNA, 0.25µl of Go Taq DNA polymerase (Promega), 1.25 µl of Green Go Taq tampon; 1,5 µl of MgCl 2 (50 mM), 0.2 mmol of each dATP, dCTP, dGTP and dTTP and 50 pmol of each primer. The PCR program consists of 30 cycles having the following parameters: denaturation at 94 °C for 1 min, hybridization at 55 °C for 1 min and annealing at 72 °C for 1 min and 30s. Amplified DNA was used as such for restriction with the enzyme HaeIII (1U) at 37°C for 3h 30.

6

PCR products were run in 2% agarose gels electrophoresis submerged in 1X Tris -borate-EDTA buffer (TBE) under 90 volts. DNA in the gel was visualized by exposure to UV light and photographed using a Gel Doc XR apparatus (Bio-rad, USA). The size of DNA fragments was estimated using a 100 bp DNA ladder (Promega).

Virulence for Fish

The virulence potential of starved cells was tested in tanks containing 50 litres of seawater maintained at 25°C with sea bass (10 g body weight) for VL13 and gilt head sea bream (10 g body weight) for VD5. Groups of ten fish were injected intraperitoneally with 0.1 ml of bacterial suspension of r esuscitated cells (10 4 -10 6 CFU/ ml). Mortality was recorded daily during a 10-day period and was considered only if the bacterium was reisolated as a pure culture from internal organs. Fish inoculated with phosphate buffer saline (PBS) and exponential phase growth cells were also included as controls. Percent mortality at the end of the experiments was recorded.

Results

1. Survival Curves during Starvation

The survival of V. alginolyticus in sterilised seawater at ambient temperature under starvation was investigated. Cultivability on MA and selective medium TCBS was studied. Our results showed that cells remained culturable for prolonged periods of time in all microcosms (Fig 1 ,   2). At room temperature (22-25°C) no difference showed in the culturability for VD4 and VD5. The survival curves fell into three stages. At the beginning of starvation, cells lost culturability rapidly within two weeks. This was followed by a slow decline in culturability and then a drop of plate counts to less than 1 cell per ml after nine months for VD4 and VD5, seven months for VL12 and six months for VL13.

The selective medium, TCBS, was compared with the non-selective medium, MA, for enumeration of culturable cells. The difference between TCBS and MA counts of strains VL12 and VL13 was about 1 log at the initial time point, but thereafter, the difference became greater. However, we did not observe any difference between TCBS and MA counts of strains VD4 and VD5. Theoretically, the plate count difference between non-selective and selective media represents the number of injured cells that have been exposed to a sublethal physical or chemical environment, such as heat, refrigeration, freezing, and so on [START_REF] Jiang | Survival of Vibrio parahaemolyticus at low temperatures under starvation conditions and subsequent resuscitation of viable, nonculturable cells[END_REF].

For the VD4 and VD5 strains, it appeared that starvation did not cause cell injury, since the difference between MA and TCBS counts was constant regardless of how long the cells were starved (Fig 1). In contrast, more cells were injured for the strains VL12 and VL13, as evidenced by an increasing difference between MA and TCBS plate counts (Fig 2 ).

Enzymatic Changes of V. alginolyticus during Starvation

The results of API 20E tests showed no modification in the biochemical profile of all V. alginolyticus strains incubated in seawater microcosms after a long period of starvation. Thus, on API ZYM galleries, and after six months of starvation, VD4 and VD5 lost many of their enzymatic activities. VD4 express only the Naphtol-AS-BI-phosphohydrolase and VD5 express the Naphtol-AS-BI-phosphohydrolase and Leucine arylamidase (Fig 3 A). In prepared mediums, VD4 lost all its virulence factors and for VD5, only caseinase, DNASE and lipase activities were detected (Table 1). However, strains isolated from sea bass (VL12 and VL13) showed a remarkable decrease in their enzymatic production level (Fig 3 B), except for the phosphatase alkaline and Naphtol-AS-BI-phosphohydrolase produced by VL13. These findings could be explained by the preparation of cells to enter into the VBNC state characterized by low metabolic rate, reduction in size, morphological changes or synthesis of specific proteins. This adaptation strategy developed by bacteria in aquatic environments allows them to survive during a long period of time.

Results of ARDRA

V. alginolyticus cells starved for as long as one year were able to regain the ability to be amplified with the addition of nutrients to starved cells from microcosms. VBNC cells were resuscitated by adding seawater microcosms to LB broth and incubation at 28°C. DNA was easily extracted and PCR amplification did not fail with the primers selected, giving a unique pattern at 1500 bp (Fig 4,lanes 12,13,14,15,16,17,18,19). ARDRA showed the same profile for the strains of V. alginolyticus isolated from gilt head sea bream (VD4 and VD5) and sea bass (VL12 and VL13). Profiles of initial cells were similar to the profile of starved cells (incubated 1year) (Fig 5,lanes 3,4,5,6,7,8,9,10). These data confirm that this was the same strain that resists in all microcosms during a long period of time.

Virulence Assay

This test was assayed for the strain VL13 and VD5. Cells from microcosms were able to infect cultured fish, gilt head sea bream (Sparus aurata) and sea bass (Dicentrarchus labrax) by i.p. injection, even after one year of starvation (Table 2). Starved cells presented a degree of virulence similar to that of non-starved cells used as controls. The retention of virulence under conditions simulating those found in the aquatic environment is additional evidence supporting the hypothesis that water is one of the natural reservoirs of this pathogen.

Discussion

Under adverse conditions, Vibrio species can survive and adapt for a long period of time, and like many other microorganisms, they can enter a viable but non-culturable state [START_REF] Morita | Bioavailability of energy and the starvation s tate[END_REF][START_REF] Oliver | The viable but nonculturable state in the human pathogen Vibrio vulnificus[END_REF]. This state was employed mainly by gram-negative bacteria and has been proposed as a strategy for survival by bacteria in the natural environment (Ben kahla -Nakbi, Besbes, Bakhrouf, 2006;[START_REF] Baffone | Retention of virulence in viable but non-culturable halophilic Vibrio spp[END_REF][START_REF] Gupte | Induction and Resuscitation of Viable but Nonculturable Salmonella enterica Serovar Typhimurium DT104[END_REF] humidity [START_REF] Heidelberg | Effect of aerosolization on culturability and viability of gram-negative bacteria[END_REF] , and aeration [START_REF] Rollins | Viable but nonculturable stage of Campylobacter jejuni and its role in survival in the natural aquatic environment[END_REF], have been found to be important factors that contribute to the induction of the VBNC state.However, the stresses of nutrient deprivation and low temperature have been shown to be the main causes of the induction of VNC stages in some pathogenic bacteria [START_REF] Oliver | Formation of viable but nonculturable cells[END_REF][START_REF] Colwell | Vibrios in the environment: viable but nonculturable Vibrio cholerae[END_REF].

This study showed the development of the VBNC state in V. alginolyticus in nutrient-limiting microcosms for prolonged periods of time. It was possible to resuscitate the non-culturable organisms by nutrient addition at 28°C. Many reports suggested that the culturable population which appears is a result of regrowth of one or more culturable cells in the population that had not been detected when the population was assayed [START_REF] Weichart | Low temperature induced nonculturability and killing of Vibrio vulnificus[END_REF][START_REF] Jiang | Survival of Vibrio parahaemolyticus at low temperatures under starvation conditions and subsequent resuscitation of viable, nonculturable cells[END_REF]. Ravel et al. (1995) reported that recovery of VNC cells of V. cholerae from starvation at 4°C in nutrient-limite d artificial seawater after a temperature upshift to 30°C was due to regrowth of surviving cells but not resuscitation of all VNC cells. However, [START_REF] Oliver | Entry into, and resuscitation from, the viable but nonculturable state by Vibrio vulnificus in an estuarine environment[END_REF] demonstrated that the rapid development of culturability represents a true resuscitation of the VBNC population to the actively metabolizing, culturable state and is not a result of growth of a few residual cells present in membrane diffusion chambers. [START_REF] Whitesides | Resuscitation of Vibrio vulnificus from the viable but nonculturable state[END_REF] noted that, regardless of the duration for which the VBNC cells were present in a high level of nutrient, the number of culturable cells ultimately present did not exceed that of control cells resuscitated at room temperature, but without being exposed to the added nutrient. They supposed that if culturable cells had been present among the VBNC cells, a population approaching 10 8 CFU/ml or greater would have been expected following these conditions of incubation. This observation argues that no culturable cells were present in the microcosms. And to further examine whether regrowth of undetected culturable cells, as opposed to true resuscitation of VBNC cells, might account for the appearance of the large culturable populations routinely observed following a temperature upshift in nutrient-free artificial seawater (ASW), the same group made different dilutions of ASW, and they concluded that resuscitation occurred in each dilution, with final culturable populations reaching levels reflecting the dilution applied.

Recently, [START_REF] Kong | Role of catalase and oxyR in the viable but nonculturable state of Vibrio vulnificus[END_REF] reported that entry of V. vuln ificus into the VBNC state is due largely to the loss of catalase activity in these cells, and that resuscitation is likely due to the production of catalase in H 2 O 2 -free diluents (e.g. ASW) at permissive (e.g. room) temperatures.

VBNC cells do indeed leave this dormant state upon removal/reversal of the inducing factor; in our case, the absence of nutrients and osmotic shock during nine months were the inducing factors of entrance into the VBNC state. Moreover, the resuscitation phenomenon occurred once these adverse conditions were reversed (addition of nutrients).

Various viability assays have been successfully employed to detect the presence of VBNC cells, such as direct viable count [START_REF] Kogure | A tentative direct microscopic method for counting living bacteria[END_REF]; florescence microscopy us ing various stains, e.g., cyanoditolyl tetrazolium chloride (CTC) ( Rodriguez, Phipps, Ishiguru & Ridgway,1992); and more recently molecular genetic methods [START_REF] Warner | Randomly amplified polymorphic DNA analysis of starved and viable but nonculturable Vibrio vulnificus cells[END_REF][START_REF] Lázaro | Viability and DNA maintenance in nonculturable spiral Campylobacter jejuni cells after long-term exposure to low temperatures[END_REF]. However, the eventual proof of vitality is through resuscitation of the cells and confirmation that the recovered cells are indeed identical to the starting cells [START_REF] Gupte | Induction and Resuscitation of Viable but Nonculturable Salmonella enterica Serovar Typhimurium DT104[END_REF].

In this study, ARDRA was used to confirm that the same strain of Vibrio alginolyticus resists in all microcosms during a long period of time and transforms into a VBNC state (Fig 5). [START_REF] Yaron | A reverse transcriptase polymerase chain reaction assay for detection of viable Escherichia coli O157:H 7: investigation of specific target genes[END_REF] reported a variety of genes, including those required for 16S rRNA synthesis, continued to be expressed in nonculturable cells of E. coli O157:H7.

However, the question posed by many authors was "Can starved cells retain their virulence factors and still remain pathogenic for humans and animals?" Long-starved cells of V. vulnificus biotype 2 maintained their ability to infect eels and humans. culturable cells from saline microcosms were able to infect eels and mice, even after two years of starvation (Macro-Noals, [START_REF] Macro -Noals | Effects of salinities and temperature on long term survival of the eel pathogen Vibrio vulnificus biotyppe 2 (seovar E)[END_REF] . Stable virulence of other bacterial species in the VBNC state was reported previously. VBNC Escherichia coli and V. cholerae cells, for example, were isolated from rabbit ileal loops in which enterotoxigenicity was exhibited (Colwell et al. 1985;Grimes & Colwell, 1986).

Vibrio spp. can be transmitted through seawater. They are attached to surfaces or surviving in sediments; they can reinfect fish farms seasonally when conditions are favourable [START_REF] Hoff | Survival of Vibrio anguillarum and Vibrio salmonicida at Different Salinities[END_REF].

Our experiences showed that, on the one hand, starved cells of V. alginolyticus can retain their virulence enzymes (VD5 and VL13) and can reinfect cultured fish even after a long period of starvation. On the other hand, starved cells of VD4 and VL12 appeared to lose all or an important part of their exoenzymes (Table 1). The pathogenicity and conservation of virulence factors seems to be a strain-specific expression. The high diversity of strains belonging to the same species may explain the strain-specific expression of virulence factors and, consequently, the different data obtained from various geographic areas [START_REF] Baffone | Occurrence and expression of virulence-related properties by environmental halophilic Vibrio spp . in in vitro and in vivo systems[END_REF]. Moreover, the loss of virulence enzymes of the two studied strains (VD4 and VL12) might be virtual, and cells could be resuscitated after passage in the animal. [START_REF] Baffone | Retention of virulence in viable but non-culturable halophilic Vibrio spp[END_REF] demonstrated that VBNC forms of two environmental strains V. alginolyticus and V. paraheamolyticus obtained by culturing bacteria in seawater microcosms and incubating at 5 °C can be resuscitated and retain their virulence properties after two consecutive passages of the strains in the rat ileal loop model.

Conclusion

In summary, our results suggest that V. alginolyticus has developed strategies that allow it to survive in seawater in the absence of nutrients and outside its natural host. Under these conditions, this bacterium maintains its potential to infect fish. Our results are in accordance with those obtained in similar studies performed with other species of Vibrio, autochthonous members of aquatic ecosystems. This means that V. alginolyticus is an opportunistic pathogen whose main reservoir is probably the aquatic environment where it can survive at different rates, depending on temperature. From these environments, the bacterium could infect susceptible fish, multiply in them following the development of vibriosis and be liberated in large numbers into the environment.

  cells (Strain before incubation in sea water microcosms), s: starved strain (after 6 months of incubation in sea water microcosms), (ß): complete haemolysis, (a ): incomplete haemolysis.

Table 1 . Exoenzymes expression (prepared mediums) by V. alginolyticus strains before and after incubation in sea water microcosms. Exoenzymes expression by strain: Exoenzymes

 1

	VD4 i	VD4 s	VD5 i	VD5 s	VL12 i	VL12 s VL13 i VL13 s

Table 2 . Mortality of fish inoculated with starved cells and of controls of V. alginolyticus during 10 days

 2 Overview of the patterns obtained after restriction of the amplified 16S rRNA gene with the enzyme HaeIII of V. alginolyticus strains. Lanes 1, 11, molecular size marker (100-bp ladder; Promega). 2, negative control. 3, 5, 7, 9 (respectively VD4, VD5, VL12 and VL13 not incubated in microcosms) 4, 6, 8, 10 (respectively VD4, VD5, VL12 and VL13 after 1 year of incubation in sea water microcosms)12, 13, 14, 15, 16, 17, 18 and 19 pattern at 1500 bp of different forms of V. alginolyticus cited above.

		Strains	Dose CFU/ml % mortality
	VL13	Control cells	6.67 x 10 4	60%
			5 x10 5	80%
		Starved cells	1.07 x 10 4	50%
			6.3 x 10 5	60%
	VD5	Control cells
			5 x10 5	60%
			4.3 x10 6	70%
		Starved cells	5 x10 5	40%
			6 x10 6	50%

1998;Gomez-Leon et al., 2005; Ben kahla-Nakbi et al., 2006). The route of infection and transmission of this species remains to be investigated. One method of transmission might be

A 24-hour culture was centrifuged at 12000g, 4°C for 15 min. The pellet was washed three times with a saline solution. Then the cells were inoculated into Erlenmeyer flasks containing

[START_REF] Whitesides | Resuscitation of Vibrio vulnificus from the viable but nonculturable state[END_REF]. Temperature(Oliver, Nilsson & Kjelleberg, 1991), as well as nutrient limitation(Fouz, Toranzo, Marco-Noales, & Amaro, 1998), levels of osmolarity,

Acknowledgment

We would like to thank Pr. JEMMALI Mongi director of the club "Young and Sciences" of Monastir for providing the local and aquaria for infection experiments.


David A Roberts
email: d.roberts@student.unsw.edu.au

Alistair G B Poore

Emma L Johnston

MBACI sampling of an episodic disturbance: stormwater effects on algal epifauna

Keywords: Stormwater, Environmental monitoring, MBACI, Algal epifauna, Pulse disturbance

Whilst it has been well established that stormwater and associated contaminants negatively impact the quality of recipient waters, the ecological effects of individual stormwater pulses in marine environments are relatively unknown. In this study, the impacts of stormwater outfalls upon water quality and epifaunal invertebrates inhabiting the alga Sargassum linearifolium are assessed through an MBACI sampling program.

Water quality and the abundance of mobile algal epifauna were recorded at three control sites and three impact sites every 10-12 weeks during dry weather and opportunistically within 24 h of large rain events (> 50 ml rainfall in 24 h) and again 4 d after the rain event. Sampling took place during two periods over two separate years and this included four large rainfall events. Following rainfall, salinity dropped rapidly at impact sites close to stormwater outfalls, whilst turbidity increased. Declines in salinity at control sites were slight and turbidity did not differ to 'before' periods. The abundance of epifaunal gastropods and polychaetes were reduced at impact sites 24 h after rain events, but not control sites. The abundance of copepods, amphipods and ostracods, however, were reduced at both control and impact sites for up to 4 d following rainfall. Reductions of these taxonomic groups could not be attributed to stormwater outfalls. Whilst short-term impacts of stormwater runoff were identified for some faunal groups, impacts were not identified for the majority. Instead, effects were harbour wide (i.e. at control and impact sites), probably in response to the physical disturbance of heavy seas associated with large rain events.

Introduction

Stormwater is a largely uncontrolled and unregulated source of contamination that has been implicated in the deterioration of water quality in freshwater [START_REF] Lieb | Effects of urban runoff from a detention pond on water quality, temperature and caged Gammarus minus (Say) (Amphipoda) in a headwater stream[END_REF] and marine systems [START_REF] Bay | Water quality impacts of stormwater discharges to Santa Monica Bay[END_REF] and in the ongoing degradation of benthic marine habitats where contaminants may accumulate [START_REF] Birch | Source of heavy metals in sediments of the Port Jackson estuary, Australia[END_REF]Taylor, 1999, Carr et al., 2000). This complex effluent may contain many toxic contaminants including heavy metals, polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs) and pesticides [START_REF] Makepeace | Urban stormwater qualitysummary of contaminant data[END_REF]. In regions such as Port Jackson, Australia, stormwater runoff is known to be one of the major sources of heavy metals [START_REF] Bickford | Aquatic ecological and human health risk assessment of chemicals in wet weather discharges in the Sydney region, New South Wales, Australia[END_REF]. There is also the potential for deleterious impacts upon marine organisms through reduced salinity around stormwater outfalls.

Laboratory tests have established the toxicity of prolonged exposures to stormwater effluent in a variety of marine and aquatic fish and invertebrates [START_REF] Carr | Impact of storm-water outfalls on sediment quality in Corpus Christi Bay, Texas, USA[END_REF][START_REF] Ellis | Ecotoxicological approaches and criteria for the assessment of urban runoff impacts receiving waters[END_REF][START_REF] Hatch | Sediment toxicity and stormwater runoff in a contaminated receiving system: consideration of different bioassays in the laboratory and field[END_REF][START_REF] Kszos | Source of toxicity in storm water: zinc from commonly used paint[END_REF][START_REF] Skinner | Developmental effects of urban storm water in medaka (Oryzias latipes) and inland silverside (Menidia beryllina)[END_REF]. Fish mortality has also been noted following periods of intense storm runoff [START_REF] Magaud | Modelling fish mortality due to urban storm run-off: Interacting effects of hypoxia and unionized ammonia[END_REF]. However, stormwater runoff events occur sporadically and may be relatively short-lived with pulses lasting from minutes to hours [START_REF] Brent | A method for the toxicity assessment of wet weather events[END_REF]Herricks, 1999, Burton Jr. et al., 2000) and the responses of communities of marine organisms to pulses of stormwater in the field remain largely unknown. Pulses of chemical contaminants and freshwater entering marine systems have the potential to cause immediate impacts localized around individual stormwater outlets [START_REF] Brent | A method for the toxicity assessment of wet weather events[END_REF]Herricks, 1999, Burton Jr., et al., 2000). Predicting the ecological impacts of pollutants upon field populations in receiving environments will be greatly enhanced by knowledge gained from ecological impact assessments [START_REF] Schroeter | Detecting the ecological effects of environmental impacts: a case study of kelp forest invertebrates[END_REF]. There have been few such studies addressing the effects of stormwater in marine systems.

The current study assessed the short-term (1-4 d) ecological impacts of stormwater events upon the mobile invertebrates that inhabit macroalgae. To date, there have been no studies that have monitored the ecological impacts of stormwater upon mobile algal epifaunal assemblages. The detection of ecological impacts in communities that are both spatially and temporally variable remains a major challenge in applied ecology [START_REF] Schroeter | Detecting the ecological effects of environmental impacts: a case study of kelp forest invertebrates[END_REF]. The unpredictable nature of episodic events, such as stormwater, poses additional difficulties (Beck, 1996, McCahon and[START_REF] Mccahon | Episodic pollution: causes, toxicological effects and ecological significance[END_REF], effectively enforcing opportunistic approaches. Nevertheless, the continued development of 'before-after, control-impact' (BACI) sampling designs has improved our ability to detect environmental impacts [START_REF] Green | Sampling design and statistical methods for environmental biologists[END_REF][START_REF] Schroeter | Detecting the ecological effects of environmental impacts: a case study of kelp forest invertebrates[END_REF].

Algal beds dominate hard substrates in temperate marine habitats and support diverse assemblages of mobile invertebrates including crustaceans, molluscs and polychaetes [START_REF] Martin-Smith | Short-term dynamics of tropical macroalgal epifauna: patterns and processes in recolonisation of Sargassum fissifolium[END_REF] which form a major component of coastal biodiversity [START_REF] Taylor | Short-term dynamics of a seaweed epifaunal assemblage[END_REF]. The epifauna of algal beds are highly mobile and posses the ability to respond rapidly to perturbations [START_REF] Martin-Smith | Short-term dynamics of tropical macroalgal epifauna: patterns and processes in recolonisation of Sargassum fissifolium[END_REF]. As such, they are a perfect candidate for assessing the ecological effects of short-lived pulse disturbances.

In the current study we examined the effects of storm water pulses through a multiple before-after-control-impact (MBACI) sampling program [START_REF] Keough | Designing environmental monitoring for pulp mills in Australia[END_REF] within the outer reaches of Port Jackson (Sydney Harbour), Australia. The MBACI sampling design was modified to incorporate multiple 'after' times in order to approximate durations of effects in potentially impacted communities. The Sydney stormwater drainage system was built during the 20 th century and consists of over 20 000 km of concrete drains and more than 200 outlets emptying to marine habitats [START_REF] Brown | Impediments to integrated urban stormwater management: the need for institutional reform[END_REF]. We asked two specific questions. First; how do water quality variables (salinity and turbidity) differ through time ('before' vs. 'after' periods) at sites receiving and not receiving stormwater inputs? Second; how does the abundance of common algal epifauna associated with the common brown alga Sargassum linearifolium differ through time ('before' vs. 'after') at sites receiving and not receiving stormwater inputs?

Biological impacts in the present study may be evidenced by a range of responses including short (< 4 d) and long-term (> 4 d) reductions in abundance at impact sites which do not occur at control sites, or slower recovery from storm related physical disturbances (e.g. heavy seas) at impact sites than control sites. We define biological impacts as changes in the abundance of organisms at impact sites from 'before' to 'after' periods that differ from changes at control sites across the same period.

Methods

Sampling design

In order to assess the impacts of stormwater upon the abundance of epifaunal communities a multi-site, multi-time monitoring program known as MBACI was run [START_REF] Downes | Monitoring ecological impacts: concepts and practice in flowing waters[END_REF]. The inclusion of multiple sites and census times is an attempt to improve upon inappropriate spatial and temporal replication from which earlier BACI designs suffered [START_REF] Downes | Monitoring ecological impacts: concepts and practice in flowing waters[END_REF][START_REF] Underwood | On Beyond BACI: sampling designs that might reliably detect environmental disturbances[END_REF].

All sampling sites were shallow rocky reefs (2-5 m in depth) containing algal communities dominated by the brown alga Sargassum linearifolium. Algal beds were approximately 20 x 20 m at all sites. Three impact sites were chosen in the outer reaches of Port Jackson including Edwards Beach, Balmoral Beach and Shark Bay (Fig 1). The incorporation of multiple impact sites reduces the possibility of confusing natural fluctuations in sampled communities with anthropogenic impacts [START_REF] Schroeter | Detecting the ecological effects of environmental impacts: a case study of kelp forest invertebrates[END_REF]. Impact sites were selected which had similar catchment activities (all impact sites receive runoff from predominately urban areas) and were directly adjacent to stormwater drains which emptied onto algal dominated rocky reefs. Three control sites were also selected, including Chinamans Beach, Rocky Point and Steel Point (Fig 1). Control sites were located in separate embayments or on different headlands from stormwater drains (at least 300 m away from storm drains) and were chosen on the basis of having similar algal communities (dominated by S. linearifolium), catchment developments (predominantly urban), aspect and degree of exposure as impact sites. The abundance and taxonomic composition of epifaunal communities at all sites were found to be similar through pilot sampling conducted in early 2004.

Sampling took place over two periods across two consecutive years; May-December in 2004 and May-September in 2005 (no rain events occurred in the later part of 2005) which coincided with the highest abundance of S. linearifolium [START_REF] Poore | Preference-performance relationships and effects of host plant choice in an herbivorous marine amphipod[END_REF]. During these periods, all sites were sampled every 8-12 weeks on randomly chosen dry days ('before' dates). Randomized sampling dates at irregular frequencies are desirable as they avoid potentially coinciding and thus confounding natural cycles in sampled communities [START_REF] Stewart-Oaten | Environmental Impact Assessment: pseudoreplication in time[END_REF]. The only constraint upon the random selection of dry sampling dates was that they were preceded by at least 24 h of zero precipitation and no substantial rainfall (maximum ~ 5 mm) occurred in the week prior to collecting 'before' samples. A total of seven 'before' dates were sampled.

All sites were also sampled within 24 h of a rain event ('after' dates). At least 50 mm of rain within a 24 h period at any of the sites constituted a rain event. Rain figures from the Australian Bureau of Meteorology website (www.bom.gov.au/hydro/flood/nsw/sydney_metro.shtml) for Mosman and Rose Bay weather stations estimated rainfall received over a 24 h period at the northern and southern sites respectively. We observed strong flows emanating from all drains when sampling 24 hr after rainfall. Drains were not observed to flow under dry sampling conditions. Sites were once again sampled 4 d after such a rain event in order to establish approximate time scales of recovery. A period of four days following sampled rain events was chosen as individuals of the alga S. linearifolium are re-colonised within one or two days of complete defaunation associated with small-scale experimental disturbances (Poore, 2005, Roberts andPoore, 2006). A total of four rain events were sampled throughout the two sampling periods.

Water quality and biological sampling

Sampling involved collecting biological samples and water quality data. At each site, 10 replicate readings of salinity and turbidity were taken in situ using a Yeo-Kal ® Model 611 Intelligent Water Quality Analyzer. Swimmers deployed the probe at haphazardly selected locations within the algal bed from which algal samples were collected whilst an assistant on shore noted the readings. At each site and on each date, 7 individuals of Sargassum linearifolium were also collected while snorkeling (selected at random from algal patches within 20 m of the stormwater outfall). Samples were transferred quickly to 1 L watertight containers with resident epifaunal invertebrates, transported to the laboratory and fixed in a 5% formalin solution. Prior to counting, algal samples were rinsed in fresh water and sieved through 300 µm mesh to retain invertebrates. Epifauna were then sorted to groups including gammarid amphipods, anemones, copepods, gastropods, isopods, ostracods and polychaetes using a dissecting microscope and finally stored in 70% ethanol. Algal wet weights were recorded and counts converted to densities (# organisms/gram algae). Algal wet weights were analysed statistically to check for any possible bias in sample size. The wet mass of algal samples collected did not differ between the levels of any factors or interactions of interest (P > 0.120 in all cases).

Data analyses

Analyses of variance (ANOVA) contrasted temporal patterns of abundance and water quality at control and impact sites. Only taxa that displayed overall mean abundances > 0.10 individuals/gram of algae at control sites during dry weather were formally analysed. These taxa included gammarid amphipods, anemones, copepods, gastropods, isopods, ostracods and polychaetes. The total abundance of all individuals collected, which included rarer taxa (<0.10 individuals/gram), was also analysed. The main interaction of interest in an MBACI design for identifying ecological impacts is the interaction between control/impact sites and before/after times [START_REF] Downes | Monitoring ecological impacts: concepts and practice in flowing waters[END_REF]. This interaction was examined using planned comparisons testing for differences in abundance between before and after sampling times at control and impact sites separately. Four comparisons were run for each variable; before rain vs. 24 h after rain and before rain vs. 4 d after rain at control sites, and again for data from impact sites.

Planned comparisons were tested over the appropriate error term taken from the full model [START_REF] Quinn | Experimental design and data analysis for biologists[END_REF].

Responses considered to be indicative of biological effects were determined a priori. Short-term biological impacts of stormwater were defined as significant reductions in abundance at impact sites 24 h, but not 4 d after rain events, in the absence of significant time effects at control sites. Long-term biological impacts of stormwater were defined as significant reductions in abundance at impact sites both 24 h and 4 d after rain events, in the absence of significant time effects at control sites. Short-term biological impacts of rain events per se were defined as significant reductions in abundance at both control and impact sites 24 h, but not 4 d after rain events whilst longterm impacts would persist for 4 days at control and impact sites. Contrasting temporal patterns of change in water quality variables at control and impact sites confirmed the appropriate choice of control locations.

Analyses of variance and planned comparisons were performed using the statistical package SYSTAT ® Version 10 (SPSS Inc.). The assumptions of normality and heterogeneity of variance were tested for each variable by examining residual histograms and scatterplots of estimates vs. residuals respectively [START_REF] Quinn | Experimental design and data analysis for biologists[END_REF]. When necessary, data were log transformed to satisfy the assumptions of ANOVA. Data for salinity were negatively skewed and required reflection prior to transformation. [START_REF] Quinn | Experimental design and data analysis for biologists[END_REF].

Results

Effects of stormwater upon water quality

Water quality parameters showed different temporal patterns at control and impact sites. Salinity was significantly lower at both control and impact sites 24 h after rainfall. Reductions in salinity were substantially greater at impact sites than control sites (reduced from 34 ppt to approximately 27 ppt and 32 ppt respectively) (Table 1, Fig. 2a).

Salinities at control and impact sites were still reduced after 4 d (Table 1, Fig 2a).

However, salinity at impact sites showed strong signs of recovery to dry weather conditions (Table 1, Fig 2a). Turbidity was significantly higher at impact sites 24 h after rainfall, than during 'before' periods (Table 1, Fig. 2b). In contrast, turbidity at control sites 24 h after rainfall did not differ from 'before' periods and was significantly lower 4 d after rainfall (Table 1, Fig. 2b).

Effects of stormwater upon epifauna

Four types of response to stormwater were identified in epifaunal taxa; short-term reductions in abundance at impact sites, delayed recovery at impact sites following rainfall, broad scale reductions in abundance across both control and impact sites and, finally, no identifiable ecological impacts. Of all taxa analysed, only the response of gastropods and polychaetes were suggestive of an impact of stormwater. The abundance of both groups was significantly lower at impact sites 24 h after rainfall, than at impact sites before rainfall (Table 2, Fig. 3d,g). In contrast, the abundance of gastropods and polychaetes after rain events at control sites did not differ significantly from abundances during dry periods (Table 2, Fig. 3d,g). After 4 d, the abundance of polychaetes had recovered to dry weather conditions whilst gastropod abundances were still substantially reduced, although not significantly so (Table 2, Fig. 3d,g). While no other taxa demonstrated significant responses to stormwater, a subtle effect was detected for copepods. The abundance of copepods was significantly lower at both control and impact sites 24 h after rainfall. However, recovery at impact sites was slower than at control sites. While copepod abundance was no longer significantly reduced at control sites after 4 d, abundances were still significantly lower than during 'before' times after 4 d at impact sites (Table 2, Fig. 3b).

The abundance of gammarid amphipods (the most numerically dominant taxa), ostracods and the overall abundance of epifaunal communities were significantly lower at both control and impact sites within 24 h of heavy rainfall than during dry conditions (Table 2, Fig. 3c,f,h). Reductions of gammarids, and the entire epifaunal community were of a greater duration than those identified for polychaetes and gastropods, with lowered abundances persisting for at least 4 d following rain events with no signs of recovery (Table 2, Fig. 3c,h). Ostracods recovered more rapidly (Table 2, Fig. 3f). The abundance of anemones and isopods were highly variable at all sampling times and were unaffected by rain events (Table 2, Fig. 3a,e).

Discussion

The detection of ecological impacts of stormwater in marine systems has proven difficult despite anecdotal evidence that toxicants such as heavy metals and hydrocarbons may have widespread impacts [START_REF] Morrisey | Factors affecting the distribution of benthic macrofauna in estuaries contaminated by urban runoff[END_REF]. The majority of studies have considered the ecological impacts of accumulated contaminants (particularly in sediments), with little consideration of effects of individual pulses of stormwater upon recipient communities. Transient influences of stormwater pulses upon various aspects of the physical environments have been identified but impacts upon biological communities may not necessarily result. For example, [START_REF] Schiff | Impacts of stormwater discharges on the nearshore benthic environment of Santa Monica Bay[END_REF] sampled infaunal habitats offshore from drainage catchments following large rain events and found stormwater to alter sediment composition and contaminant concentrations in benthic habitats. However, infaunal assemblages sampled from sites within impacted regions were both abundant and healthy [START_REF] Schiff | Impacts of stormwater discharges on the nearshore benthic environment of Santa Monica Bay[END_REF].

In the present study, whilst stormwater was found to affect water quality variables, biological effects of runoff were relatively subtle and for the majority of taxa, could not be detected. Rather, the data presented in the current study show obvious and substantial impacts of actual rain events which occur across all sampling sites. Salinity at impact sites decreased substantially within 24 h of rainfall, whilst turbidity increased.

Salinity dropped slightly at control sites after rain, and turbidity did not differ from the 'before' period. In contrast to patterns in water quality, stormwater did not have consistent effects upon the abundance of epifauna, with a variety of responses identified.

Gastropods and polychaetes were the only taxa for which localized pulses of stormwater were found to have negative ecological impacts. Both taxa recovered very rapidly from pulses with the detectable impacts identified 24 h after strong rain events, no longer evident after 4 d. Such a rapid recovery suggests that reductions in the abundance of epifauna were due to emigration of mobile organisms away from stormwater plumes, rather than direct mortality. Further research examining the role of behavioural avoidance of stormwater pulses in the field by mobile epifauna is required. The speed of recovery also suggests that impacts were confined to a relatively small region surrounding stormwater drains (the current study sampled a region of approximately 20 x 20 m around stormwater drains). Thus it appears that ecological impacts of stormwater pulses in this part of the harbour have limited spatial and temporal extent.

Despite detectable impacts upon gastropods and polychaetes, temporal fluctuations in the abundance of the epifaunal community as a whole could not be attributed to stormwater as the responses of epifauna to rain events were comparable at both control and impact sites. Although water quality parameters were clearly impacted by stormwater, the majority of taxa demonstrated either no response to stormwater pulses or significant reductions at both control and impact sites. Together the water quality and biological data suggest that whilst some ecological impacts of stormwater pulses are evident, generally, the factors that determine epifaunal abundance in algal habitats were operating across broader spatial scales than the factors that determine local water quality following heavy rainfall. It is possible that deleterious environmental conditions associated with large rain events masked any direct effects that stormwater may have upon epifaunal communities. All rain events sampled generally coincided with heavy seas, whilst conditions during dry weather sampling were typically calm.

Mobile marine assemblages have received much less attention than their sessile counterparts with respect to the role of disturbance in structuring communities [START_REF] Sousa | Natural disturbance and the dynamics of marine benthic communities[END_REF]. There is however, some evidence to suggest that the mobile epifauna of algal beds are negatively affected by turbulent seas. Night time sampling of dispersing amphipods in algal beds found greater numbers in the water column during heavy seas than calm conditions, suggesting dislodgement of amphipods from seaweeds by strong currents and large waves [START_REF] Fincham | Periodic swimming behaviour of amphipods in Wellington Harbour[END_REF]. In addition, the abundance of epifauna is frequently lower on algae at exposed sites than the same algal habitat in protected areas [START_REF] Sánchez-Moyano | Effect of environmental factors on the spatial distribution of the epifauna of the alga Halopteris scoparia in Algeciras Bay, Southern Spain[END_REF][START_REF] Tararam | The mobile fauna -especially Gammaridea -of Sargassum cymosum[END_REF][START_REF] Yassini | Ostracod fauna associated with sublittoral kelp forest vegetation at Windang Island, New South Wales, Australia[END_REF]. It is likely therefore, that the reductions identified in epifaunal abundance after rainfall can in part be attributed to heavy seas which co-occur with large storms.

The ability of an MBACI design to detect anthropogenic impacts relies in part upon the appropriate selection of control sites. Control sites must be selected that are located far enough away from the putative contaminant sources so as to be unaffected by it [START_REF] Downes | Monitoring ecological impacts: concepts and practice in flowing waters[END_REF]. In the present study, control sites showed evidence of ecological impact however, water quality data suggested that the location of control sites was appropriate for this study. Previous research has found salinity to be the most accurate means of identifying plumes of stormwater [START_REF] Washburn | Spatial scales and evolution of stormwater plumes in Santa Monica Bay[END_REF]. The combined salinity and turbidity measurements recorded in the current study suggest that stormwater plumes did not extend to control sites. This study therefore highlights the dual considerations when selecting and locating appropriate control sites. Reference locations in monitoring studies must be located outside the zone of potential environmental impact, but also close enough to impacted sites so as to be similarly affected by broader scale processes that would influence the temporal patterns of abundance in monitored communities. For example, in the present study control sites were located such that they were outside the influence of stormwater plumes, but in a region that would be similarly affected by physical disturbances associated with storms. Had control sites been located further afield (e.g. in an adjacent estuary) the impact of stormwater drains is likely to have been overestimated.

In the present study, we have attempted to approximate time scales of recovery of epifaunal communities from large rain events. Recovery from disturbance in algal epifaunal communities involves re-colonisation by dispersing epifauna in the water column or organisms crawling from alga to alga [START_REF] Taylor | Short-term dynamics of a seaweed epifaunal assemblage[END_REF] and can be very rapid, with recovery from complete defaunation within days [START_REF] Poore | Scales of dispersal among hosts in an herbivorous marine amphipod[END_REF]. Epifaunal communities as a whole were reduced throughout the outer reaches of a large harbour in the present study. In a system where recovery is reliant upon re-colonisation of small dispersers, it is likely that the speed at which communities recover is dependant in part upon the spatial scale of disturbance. Recovery from a disturbance event on a harbourwide scale is likely to be much slower than recovery from localized disturbances. This appears to hold true for the dominant fauna in our study (amphipods and copepods); abundances of which were still depressed four days after a rainfall event.

The impacts of stormwater pulses on algal epifauna in this study appear to be subtle and quite localized. Contaminants within stormwater may pose a greater threat to marine organisms through their tendency to accumulate over time [START_REF] Bay | Water quality impacts of stormwater discharges to Santa Monica Bay[END_REF][START_REF] Carr | Impact of storm-water outfalls on sediment quality in Corpus Christi Bay, Texas, USA[END_REF]. Experimentally increased copper levels in the tissues of the alga S. linearifolium results in reduced colonisation, feeding and survivorship of Peramphithoe parmerong, an abundant herbivorous amphipod (Roberts et al., 2006). Consequently, the ecological impacts of contaminants in surface runoff may become more evident when sampled across longer time scales than individual rain pulses. The impact of individual pulses of contaminants may also decrease with time, as background levels of contamination increase, resulting in constant stresses upon exposed communities, and subsequent development of resistance [START_REF] Bryan | Adaptation of the polychaete Nereis diversicolor to estuarine sediments containing high concentrations of heavy metals. Part 1 General observations and adaptation to copper[END_REF]Hummerstone, 1971, Shavyrina et al., 2001). Further research is required in order to link the findings of laboratory based stormwater toxicity tests with the results obtained from field monitoring programs. Means marked with circles indicate a significant difference exists between the impact site mean at that time and the impact site data from before periods. gastropods, e) isopods, f) ostracods, g) polychaetes and h) the total community at control (solid line) and impact sites (dashed line). Data are mean densities ± standard error (n = 3 per observation, means of site averages). Planned comparisons contrasted the abundance of taxonomic groups at control and impact sites separately. Means marked with triangles indicate a significant difference exists between the control site mean at that time and the control site data from before periods. Means marked with circles indicate a significant difference exists between the impact site mean at that time and the impact site data from before periods.

Figure 1 .

 1 Figure 1. Location of storm water drains (open stars), control (open circles) and impact (closed

Figure 2 .

 2 Figure 2. Temporal changes in a) salinity and b) turbidity at control (solid line) and impact sites

Figure 3 .

 3 Figure 3. Temporal changes in the abundance of a) anemones, b) copepods, c) gammarids, d)

Table 1 .

 1 Planned comparisons for water quality parameters contrasting temporal changes at control and impact sites. Appropriate error terms were taken from the omnibus ANOVA.Data are ppt (salinity) and nephelometric turbidity units (turbidity). Bold figures are statistically significant (P < 0.05).

			Salinity b	Turbidity a
		df	MS	p	MS	p
	Control sites
	Before vs. 24 h After 1	4.003	0.006	3.382	0.389
	Before vs. 4 d After 1	6.986	0.001 26.619 0.034
	Impact sites
	Before vs. 24 h After 1 35.557 0.000 24.386 0.040
	Before vs. 4 d After 1	7.720	0.001 19.364 0.061
	Error	8	0.282		4.073

a Log transformed.

b Reflected and log transformed.

Table 2 .

 2 Planned comparisons contrasting temporal changes in the abundance of epifaunal taxa at control and impact sites. Appropriate error terms were taken from the omnibus ANOVA.

			Anemones a	Copepods a	Gammarids a	Gastropods a
		df	MS	p	MS	p	MS	p	MS	P
	Control sites
	Before vs. 24 h After 1	2.345	0.169 14.949 0.043	4.293	0.030	2.715	0.298
	Before vs. 4 d After 1	0.642	0.452	7.863	0.120 13.034 0.002	0.864	0.548
	Impact sites
	Before vs. 24 h After 1	0.209	0.664 70.324 0.001 21.713 0.000 19.003 0.019
	Before vs. 4 d After 1	0.131	0.730 17.208 0.033	7.732	0.008	6.020	0.136
	Error	8	1.027		2.599		0.623		2.195
			Isopods a	Ostracods a	Polychaetes a	Total a
		df	MS	p	MS	p	MS	p	MS	p
	Control sites
	Before vs. 24 h After 1	0.053	0.674 11.050 0.017	0.032	0.894	7.647	0.024
	Before vs. 4 d After 1	0.272	0.350	0.437	0.567	0.541	0.586	7.438	0.026
	Impact sites
	Before vs. 24 h After 1	0.000	0.995	9.072	0.026 23.844 0.006 33.414 0.000
	Before vs. 4 d After 1	0.238	0.380	0.358	0.603	1.241	0.416	8.532	0.019
	Error	8	0.276		1.224		1.684		0.995
	Data are mean abundances (# of individuals/gram of algae). Bold figures are statistically significant (P <
	0.05).

a Log transformed

Acknowledgements

The authors wish to acknowledge the assistance of Nicole Hill, Graeme Clark, Candida Barclay, Bronwyn Cumbo and Karen Lee with field work for this research. We thank Richard Piola, Jan Carey, Robert Spies and three anonymous reviewers for their comments which have greatly improved the manuscript. All research was conducted under New South Wales Fisheries permit # F85/122(C). Research was supported by a UNSW Faculty Research Grant awarded to AGBP and ELJ and an Australian Research Council Discovery Project awarded to AGBP.