Interactions between Zn and bacteria in marine tropical coastal sediments

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    Interactions between Zn and bacteria in marine tropicalcoastal sediments

    Olivier Pringault & Hlna Viret & Robert Duran

    Received: 6 July 2011 /Accepted: 11 September 2011 /Published online: 28 September 2011# Springer-Verlag 2011

    AbstractPurpose The main goals of this study were (1) to examinethe effects of zinc on the microbial community structure ofanthropogenically impacted sediments in a tropical coastalecosystem and (2) to determine whether these microbialbenthic communities may enhance the adsorption of zinc.Methods The interactions between zinc and bacteria intropical sediments were studied in sediment microcosmsamended with 2.5 mg L1 of Zn in the water phase andincubated for 8 days under different environmental con-ditions, oxic/anoxic and glucose addition. At the end ofincubation, microbial structure was assessed by molecularfingerprints (T-RFLP) analysis and Zn speciation in thesediment was determined by sequential extraction.Results In the three studied sediments, Zn spiking resulted inonly slight changes in bacterial community structure. Incontrast, the addition of low concentrations of glucose (5 mM)strongly modified the bacterial community structure:

  • capacities, can remobilize metal from sediments (Flemminget al. 1990; Petersen et al. 1997). This ability to mitigate thefate of metals opens up interesting perspectives for the usethese microorganisms in bioremediation applications (Gadd2000).

    Coastal sediments support many different kinds ofbacterial metabolism involving degradation of organicmatter under both oxic and anoxic conditions. Consequent-ly, benthic bacterial communities can influence the fate of ametal in various ways, employing a number of differentmetabolic processes. For example, sulfate-reducing bacte-ria, through the production of sulfide, which strongly bindsdivalent heavy metals, could play a major part in metalremoval at the sedimentwater interface if the environmen-tal conditions are conducive to sulfate reduction (White etal. 1997). As a consequence, benthic bacterial communitiesmay directly interact with the fate of metal by, e.g., sulfideproduction or indirectly interact by oxygen depletion thusfavoring sulfide production. As in most coastal environ-ments, the inshore waters of New Caledonia are subject tolarge increases in urbanization together with an importantdevelopment in nickel mining activities. Unfortunately, thisdevelopment is concomitant with inadequate wastewatertreatment, resulting in large inputs of heavy metals such asZn, Cr, and Ni into the lagoon (Fernandez et al. 2006).Coastal sediments in the vicinity of Nouma (NewCaledonia capital, 100,000 inhabitants in 2000) exhibithigh levels of Zn, Ni, and Cr contamination (Dalto et al.2006). Such conditions are, therefore, favorable for theestablishment of microbial communities that may tolerateelevated metal inputs. Indeed, metal supplies can structurebenthic microbial communities by exerting a selectionpressure on nonmetal-resistant bacteria in favor of metal-tolerant bacteria (Mertens et al. 2006; Nazaret et al. 2003;Rasmussen and Srensen 2001).

    In this study, special interest has been given to Znbecause high levels (525 g/L) are often observed in thewater column in the vicinity of Nouma due to water runoffof the surrounding area. This Zn contamination stronglymodified the phytoplankton community structure in pristineareas, whereas in polluted sites, the effects were lesspronounced (Rochelle-Newall et al. 2008). Similarly, recentstudies have shown that benthic microbial communities inNew Caledonia sediments can maintain their metabolicactivity under Zn concentrations that are known to beinhibitory for nonmetal-tolerant bacteria (Pringault et al.2008; Viret et al. 2006). To understand fully the impact ofmetal on microorganisms, the consequences of metalpollution on microbial diversity must also be assessed inorder to determine the potential role of the pollutants onmicrobial community structure. The pollution-inducedcommunity tolerance (PICT) concept (Blanck 2002) sug-gests that the microbial community structure of polluted

    sites will be only slightly modified upon contaminantexposure due to pollution adaptation, as it has beenobserved for phytoplankton communities upon Znexposure (Rochelle-Newall et al. 2008) or for pesticides(Pesce et al. 2009; Vercraene-Eairmal et al. 2010) orpolycyclic aromatic hydrocarbon (Lekunberri et al. 2010).However, the link between PICT and microbial structure isnot always straightforward (Brard and Benninghoff2001) and nonpollution-tolerant populations may domi-nate in environments where species adapted to contami-nation are expected to dominate (Pringault et al. 2008).Environmental factors (nutrient availability, irradianceconditions) can strongly influence the selective actionof the pollutants on the microbial community structure,depending on the season and the microbial succession(Brard et al. 1999; Guasch et al. 1997; Lekunberri et al.2010). In addition, the effects of a metal on microbialdiversity will also depend on its bioavailability, which canbe modified by the metabolic processes involved in its fate(Gillan 2004; Gillan et al. 2005). In other words, metalconcentrations can be high in a coastal environment, but ifit is present in an innocuous form, it is likely that it wouldexert only a weak pressure on microbial communityselection.

    The main goals of this study were (1) to examine theeffects of zinc on the microbial community structure ofanthropogenically impacted sediments in a tropicalcoastal ecosystem and (2) to determine whether thesemicrobial benthic communities may enhance the adsorp-tion of zinc. For that purpose, microcosms were set upunder different environmental conditions in order toinvestigate some of the main factors that might controlthe fate of Zn and its interactions with the microbialcommunity.

    2 Material and methods

    2.1 Sampling procedure

    Sediment was collected with a Van Veen grab sampler(Hydrobios). The top surface (approximately 1 cm thick)was collected with a sterile plastic spoon and immediatelystored at ambient temperature (in degrees Celsius) anddarkness in a plastic box tightly sealed to avoid aircontamination. Seawater was sampled with a Niskin bottleat approximately 1 m above the sediment surface. Threedifferent stations were sampled (Fig. 1). The first stationwas located close to the city of Nouma (St. Marie Bay[SM]) The coast around this station is highly urbanizedand at the head of the bay there is a sewage outfall thatdeposits minimally treated household and urban wasteand runoff to the bay. The second station was located

    880 Environ Sci Pollut Res (2012) 19:879892

  • close to the mouth of the Dumba River (Dumba Bay[DB]) and is subjected to both anthropogenic andterrigenous inputs. The latter was located in a smallbay 20 km north of Nouma and is subjected toagrochemical inputs (Port Laguerre [PL]). A moredetailed description of the sampling areas can be foundin Ouillon et al. (2010) and Fichez et al. (2010). Sedimentcharacteristics (granulometry, carbonate content, and or-ganic carbon content) were analyzed according to theprocedures described in Fernandez et al. (2006).

    2.2 Microcosm setup

    Upon return to the laboratory (

  • sterile N2 for half an hour and then the microcosms weretightly closed. Microcosms were labeled as follows: B+ or B

    for microcosms nonsterilized or sterilized, respectively, andM+ or M for microcosms amended with Zn2+ or non-amended with Zn2+, respectively. Glu designates the micro-cosms amended with glucose. For each oxic/anoxiccondition, one microcosm (B+M) was incubated withoutmetal spiking as a control, one microcosm (B+M+) wasamended with 2.5 mg L1 of zinc (ZnCl2), and onemicrocosm (BM+) was amended with zinc (2.5 mg L1)and formaldehyde (43 ml final volume) in order to sterilizethe sediment and the water column. A fourth microcosm(B+M+Glu) was amended with zinc (2.5 mg L1) andglucose (5 mM, final concentration) in order to stimulatebacterial metabolic processes as suggested by Diaz-Ravinaand Baath (1996). A fifth microcosm (B+MGlu) was alsoincubated with glucose and bacteria but without Zn in bothoxic and anoxic conditions. Microcosms were incubatedduring 8 days and samples were collected daily for theanalysis of metal concentration in the water phase. Samplingof the water phase was performed under sterile conditionsand under a permanent flux of sterile N2 for the microcosmsincubated under anoxic conditions.

    2.3 Metal analysis

    Dissolved concentration of metal in the water phase wasmonitored daily during the microcosm incubation. For thatpurpose, a volume of 10 ml (in triplicate) was sampled andfiltered through 0.45 m pore membrane (GF/C). Acidifi-cation of the samples was achieved with nitric acid (3%final concentration). Dissolved concentration of metal wasanalyzed using an inductively coupled plasma opticalemission spectrometry (ICP-OES) following the methodof Moreton et al. (2009). At the beginning and at the end ofincubation, repartition of the metal within the different solidphases of the sediment was determined using the sequentialextraction according to the procedure of Rauret et al. (1999).This sequential extraction allows to separate three distinctfractions, phase a (exchangeable, acid- and water-solublefraction), phase b (reducible fraction), and phase c (oxidiz-able fraction). Phase a corresponds to the most bioavailableform, phase b corresponds to the form associated with ironand manganese oxyhydroxides, and phase c corresponds tothe form associated with organic matter. Zn concentrations ineach phase were then measured using an ICP-OES followingthe method of Moreton et al. (2009).

    2.4 Nutrient analysis

    Nutrient concentration was measured in the watercolumn. Ammonium concentration was measured fluoro-metrically, according to the protocol of Holmes et al.

    (1999). The concentration of nutrients (NO2+NO3, PO4,and SiO3) was measured using an Autoanalyzer III(Bran+Luebbe), according to Raimbault et al. (1990).More details of the analytical protocols can be found inGrenz et al. (2010)

    2.5 Microbial population analysis

    The microbial community structure of the sediments wasanalyzed at the beginning and at the end of incubation.Sediment was homogenized prior to distribution into 2 mlEppendorf tubes. The tubes were immediately stored at80C until DNA extraction. Total DNA was extractedfrom the sediment using the UltraClean Soil DNA IsolationKit using the alternative lysis method (MoBio LaboratoriesInc., USA). The extracted genomic DNA samples werestored at 20C until further processing. Bacterial smallsubunit of ribosomal RNA (16S rRNA) genes wereamplified by polymerase chain reaction (PCR) using theuniversal primers pair R1nF and U2 (annealing to con-served regions of bacterial 16S rRNA) that amplify afragment of approximately 1,060 bp of the 16S rRNA gene.The fo rwa rd p r ime r R1nF (5 -GCTCAGATTGAACGCTGGCG-3) corresponded to positions 22 to 41of Escherichia coli 16S rRNA, and the reverse primer U2(5-ACATTTCACAACACGAGCTG-3) corresponded tothe complement of positions 1,085 to 1,066. Amplificationwas performed according to the procedure described byFourans et al. (2004). As suggested by Osborn et al.(2000), we used the terminal restriction fragment at the 5end (5-T-RF) to obtain a greater polymorphism. PrimerR1nF was fluorescently labeled with 5-tetrachloro-fluorescein. The PCR products were purified using GFXcolumns (cutoff, 100 bp; GE Healthcare, USA) prior todigestion. The restriction enzyme used in terminal restric-tion fragment length polymorphism (T-RFLP) analysis wasHaeIII (New England Biolabs, Beverly, UK). Fragmentanalysis was performed with an ABI Prism 310 (AppliedBiosystems, Foster City, USA). Dominant T-RFs from 35 to500 bp greater than 30 fluorescent units in intensity wereselected. T-RFLP profiles were normalized by calculatingrelative abundances from fluorescence intensity of each T-RF, according to Edlund and Jansson (2006). Whileassignment of identities may be uncertain using T-RFLPanalysis (Osborne et al. 2006), it does not preclude the useof this technique to compare whole communities. Profilesgenerated from different samples can be compared to assessthe similarity between communities (Whittaker index),allowing spatial or temporal changes to be detected withoutthe necessity of identifying each peak in the profiles(Danovaro et al. 2006; Pringault et al. 2008). This approachhas been largely used to compare the effects of physical andbiotic factors on sediment microbial communities (Bordenave

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  • et al. 2007; Edlund et al. 2006; Pereira et al. 2006) usingmultivariate statistical analysis (Dollhopf et al. 2001;Pringault et al. 2008).

    2.6 Mathematical analysis

    The concentration of dissolved metal in the overlying waterof the sediment can be described by the Langmuirs law(Schmitt et al. 2001):

    rt k Ct qe qt 1where r represents the adsorption rate, k the constant ofadsorption, q(t) the quantity of metal adsorbed onto thesediment, and qe the quantity of adsorbed metal at theequilibrium. q(t) can be calculated from the mass balanceequation:

    qt Co Ct V=m 2where Co represents the initial concentration of metal, C(t)the concentration of metal as a function of time t, V thevolume of the water phase (800 ml), and m the mass of thesediment (80 g). At the end of incubation, the final quantityof adsorbed metal (qf) was calculated and compared withthe estimated qe.

    In order to estimate the similarity between two T-RFLPprofiles, the Whittaker similarity index (W) was calculatedusing the following equation:

    W 1Xn


    ai1 ai2j j2


    where ai1 and ai2 are the percentage contributions toamplified DNA of the ith T-RF in samples 1 and 2,

    respectively. Since this index takes into account T-RFsrelative abundances, it provides a better estimate of thesimilarity between two microbial communities (Hewsonand Fuhrman 2006).

    Variations of microbial community structure wereassessed by correspondence analysis (CA) performed onall data from T-RFLP profiles according to the proceduredescribed by Fourans et al. (2006) and Duran et al. (2008).CA was performed with MVSP v3.12d software (KovachComputing Service, Anglesey, Wales). In all calculations,we assumed, as explained by Luna et al. (2004), that thenumber of T-RFs represents the species number and thattheir relative peak height represents the relative abundanceof each component. Relative abundances of T-RFs havebeen transformed with arcsine (x0.5), according to Legendreand Legendre (1998), to get a normal distribution of thedata since it is a condition required before applyingmultivariate statistical analysis (Dollhopf et al. 2001).

    3 Results and discussion

    3.1 Sediment composition

    The three stations had a similar sediment compositionregarding mud content (Table 1). Mud was alwaysdominant and the proportion of mud (fraction 90% in the three stations. Differences wereobserved for the carbonate fraction, with values up to61.4% for SM, and lower values for PL and DB sediments,with 18.7% and 48.7% carbonates, respectively. All threesediments exhibited similar organic carbon content (12%).Zn concentration was in the same order of magnitude forthe three studied sites and coherent with previous studies(Dalto et al. 2006, Fernandez et al. 2006).

    3.2 Bacterial community structure and impact of Zn

    Bacterial community structure was assessed by T-RFLPanalysis. The number of T-RFs varied between 12 and 34(Table 3), which is consistent with the number of ribotypesobserved in other coastal benthic environments (Luna et al.2004; Pereira et al. 2006; Urakawa et al. 2001). The three

    Table 2 Percentage of similarity (Whittaker index, W) of the initialmicrobial composition analyzed by T-RFLP for the three sediments

    Sediment PortLaguerre

    Bay ofSt. Marie

    Bay ofDumba

    Port Laguerre 100

    Bay of St. Marie 54.7 100

    Bay of Dumba 53.2 62.5 100

    Table 1 Sediment composition of the three stations (see Fig. 1 for localization)

    Station Fraction

  • stations had different communities with only 5262% ofsimilarity (Table 2). Nevertheless, microcosm incubationinduced structural changes that were different according tothe sediment origin and incubation conditions. The impactof Zn on bacterial structure was analyzed by CA (Fig. 2)and by comparing the percentage of similarity of themicrobial structure with the initial situation (Table 2). Forall sediments, the two axes explained more than 50% of theobserved structure (72.2%, 58%, and 85.4% for PL, SM,and DB sediments, respectively). The most pronounceddifferences between initial and final microbial structurewere observed for the microcosms amended with glucose(around 12.931.9% of similarity according to sedimentorigin; Table 3). For SM and PL sediments, all T-RFLPprofiles were grouped in the same cluster (Fig. 2a, b),

    suggesting that the impact of Zn moderately affected theinitial microbial structure, regardless of the oxic conditions(average Whittaker index of 6911%). Interestingly, whenall sediments spiked with Zn were included in the CA,those subject to a glucose addition were grouped in thesame cluster irrespective of sediment origin (Fig. 2d). Thiscluster was mainly structured by the presence of a commonT-RF, T-RF 179, that was the most dominant T-RF withrelative abundances of up to 78% (DB sediment, AnoxicB+M+Glu). According to the database of the ribosomalproject (, this T-RF might beattributed to Marinobacterium jannashii (Satomi et al.2002). Previous studies have noted the ability of thisaerobic heterotrophic bacterium to grow on copper (Little etal. 1996). The second cluster regrouped all the PL

    Fig. 2 CA of the microbial communities as a function of the microcosm treatments. a Port Laguerre (PL), b St. Marie Bay (SM), c Dumba Bay (DB),d compiled analysis including the microcosms amended with glucose and Zn

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  • sediments, and the SM and DB sediments were grouped ina third cluster. The average Whittaker index calculated withZn microcosms (6911%) was significantly lower(t test, p
  • the same sediment (225 versus 790 g Zn h1 in DBsediment, Oxic B+, and Anoxic B, respectively; Fig. 3f).The highest final concentration of residual zinc wasobserved in PL sediment with almost 50% (1,124 g L1)for Anoxic B and 22% (1,124 g L1) for Oxic B of theinitial concentration remaining in the sterilized microcosms.Remarkably, in all sediments, addition of 5 mM of glucosestrongly enhanced the adsorption of Zn, leading to a totaldisappearance of the metal in the dissolved phase in themicrocosms, regardless of the oxic conditions. In all of themicrocosms amended with glucose, a sharp decrease of Znconcentration was observed after few days of incubation(Fig. 3a, c, e). The observed trend of the Zn concentration

    time course did not follow Langmuirs model. Using theconcentration measured at the end of incubation, theamount of Zn bound on to the sediment (qf) can becalculated as a function of the microcosm incubationconditions. For the three sediments, adsorption of Zn wasgreater in the nonsterilized microcosms. This indicates thatthe presence of living (active) bacteria significantly en-hanced Zn adsorption. However, it should be kept in mindthat dead bacteria can also contribute to metal adsorption(Lopez et al. 2002; Selatnia et al. 2004), with ratessometimes similar to those measured with intact cells(Lopez et al. 2002). In the present study, the differencesobserved between nonsterilized and sterilized microcosms

    Fig. 3 Left panels time courseof Zn concentration in the waterphase as a function of themicrocosm treatments. Solidlines represent the fitting of theLangmuirs law (Eq. 2, see textfor details). For the microcosmsamended with glucose (Glu+),the observed kinetics did notfollow the Langmuirs law;therefore, no fitting wasperformed. Right panelsAbsorption rate (r) as a functionof time. Sediment from PortLaguerre (a, b), sediment fromSt. Marie Bay (c, d), and sedi-ment from Dumba Bay (e, f).B bacteria, M zinc, Glu glucose

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  • Fig. 4 Zn speciation (BCRdetermination) in the sedimentas a function of the microcosmtreatments at the end of incuba-tion and in the initial sedimentsampling. B bacteria, M zinc,Glu glucose. a Sediment ofPort Laguerre, b sediment ofSt. Marie Bay, c sediment ofDumba Bay. See text for adetailed description of the BCRprotocol

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  • in terms of the amount of Zn adsorbed (qf) clearly indicatethat bacterial communities do actively participate in Znbiosorption, participation that is strongly enhanced whenglucose is added as an external carbon source. At the end ofincubation, all glucose microcosms smelled strongly ofsulfide, suggesting that the addition of glucose led tosulfide production by sulfate-reducing bacteria, a com-pound which has a very high affinity for divalent metals(Fortin et al. 1994; Yu et al. 2001). Sulfate reductionmetabolism is an important metabolic process in coastalmarine environments as it can contribute up to 50% oforganic matter mineralization (Jrgensen 1977). As men-tioned above, sediments of New Caledonia are not rich inlabile carbon (Clavier et al. 1995; Clavier and Garrigue1999) and, as a consequence, sulfate reduction is carbon-limited (quality and quantity) and organic matter ispreferentially mineralized by aerobic respiration (Froelichet al. 1979). The addition of glucose stimulated sulfideproduction in two ways. Directly, through the supply oflabile carbon molecules to sulfate-reducing bacteria and,indirectly, through the stimulation of aerobic respiration, whichincreases the oxygen demand thus leading to anoxic situations,conditions that are favorable for the growth of sulfate-reducingbacteria. Nevertheless, although stimulation of sulfate produc-tion by carbon supply can enhance metal adsorption ontosediments, the bioavailability of the metal for the microbiotaand, by extension, to the whole ecosystem will strongly dependon the stability of the sediment surface. Metals bound to acidvolatile sulfides (AVS) can be easily released in the watercolumn by AVS oxidation after sediment resuspension(Canavan et al. 2007; Fang et al. 2005; Simpson et al.1998) induced by strong storms or typhoons, meteorolog-ical events that are common in tropical environments.

    Zn2+ adsorption kinetics showed that bacteria activelyparticipate in the removal of Zn from the water columnthrough the sediment, thus favoring a decrease in metalbioavailability. Nevertheless, zinc bioavailability is alsocontrolled by the Zn repartition in the sediment.

    3.4 Zn sediment repartition

    The Zn repartition in the different solid phases of thesediment was measured at the beginning and at the end ofincubation (Fig. 4). Before spiking, the total amount ofZn2+ was similar in SM and DB sediments with

  • Fig. 5 Nutrient flux in thenonsterilized microcosms(calculated from the differencebetween the final (Cfinal) and theinitial (Cinit) concentrationsmeasured in the water phase) atthe sedimentwater interface asa function of the microcosmtreatments. Averagestandarddeviation

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  • metabolism as a function of microcosm incubation conditionseven so no metabolic activity was directly measured. Incontrol microcosm (without metal spiking), flux of ammoni-um were slightly negative (approximately 3 nmol cm2) andsimilar to the rates observed in situ (Grenz et al. 2010). Underanoxic conditions for all sediments and under oxic conditionsfor PL sediment, Zn spiking resulted to an inversion of theflux (with values of up to 220 nmol cm2, for PL in anoxicconditions), indicating that the sediment became a net sourceof ammonium for the water column. Interestingly, whenglucose was added, production of ammonium was lessimportant, indicating that the stimulation of bacterial processwas concomitant with a stimulation of the ammonium uptakeby bacteria, especially in anoxic conditions. For nitrate, theimpact of Zn was less evident with both stimulation andinhibition as a function of the oxic conditions in PL sediment.In oxic and anoxic conditions, addition of glucose resulted ina strong decrease of the nitrates fluxes, suggesting thus, asobserved for ammonium, that enhancement of bacterialmetabolic processes have also stimulated the uptake ofnitrate, leading to a decrease of the flux through the watercolumn. As a general rule, phosphate and silicate fluxes wereboth stimulated by the addition of Zn regardless of thesediment origin and the oxic conditions. A higher stimulationof phosphate and silicate fluxes was the most often observedwith glucose, but the stimulation could be even higher whenboth Zn and glucose were added to the microcosm.Stimulation of nutrient fluxes by highly labile organic matterhas been observed in microcosms simulating stormwatersediment deposit (Nogaro et al. 2007). You et al. (2009)observed an inhibition of ammonia and nitrate uptake ratesby bacteria upon metal spiking in activated sludge.Similarly, nitrification (ammonium and nitrite oxidation)can be inhibited by Zn with EC50 (Zn concentrationcorresponding to 50% inhibition) in the same range as thatused in this study (in milligrams per liter) (Hu et al. 2004;Juliastuti et al. 2003).This could explain the strongaccumulation of ammonium in the water column uponZn spiking. The impact of Zn on nitrate metabolism is lessevident, which is consistent with the study of Magalhaeset al. (2007) that observed both stimulation and inhibitionof denitrification upon Zn spiking as a function of thesediment type and the environmental conditions. Althoughthe extrapolation of these results to natural environmentshas to be done with caution, overall, these results showthat Zn spiking impacted nutrient cycles in the sedimentby affecting microbial metabolisms, leading to an en-hancement of the nutrient supplies through the watercolumn. Consequently, Zn spiking can indirectly impactthe functioning of the water column by stimulatingphytoplankton and bacterioplankton that are naturallynutrient-limited in this oligotrophic ecosystem, thusfavoring eutrophication of the environment.

    4 Conclusion

    This study shows that Zn spiking in polluted sediments hasa moderate impact on microbial diversity, thus suggesting arelative tolerance of the microbial community. Neverthe-less, the observed tolerance might also be due to the highrates of bacterially mediated metal adsorption onto sedi-ments and to its transformation into an innocuous form,thus decreasing its bioavailability and its potential toxiceffect. While microbial diversity remained unaffected oronly slightly affected by Zn spiking, nutrient cyclingexhibited significant changes with an enhancement ofnutrient supplies to the water column. This clearly showsthat, even if Zn has little effect on microbial diversity inpolluted systems, the observed changes in microbialmetabolisms faced with Zn spiking can have strongconsequences in the functioning of the ecosystem byincreasing eutrophication. Further works are required toassess the mechanisms of action of Zn at the bacterial celllevel in order to better understand the impact of Zn spikingon the nutrient uptake by bacteria and its consequences forthe ecosystem.

    Acknowledgements This work was supported by the ProgrammeNational Environnement Ctier (PNEC), Chantier Nouvelle Caldonieand by the Ministre de lOutre Mer. Dr. Emma Rochelle-Newall isgratefully acknowledged for her helpful criticisms on an early version ofthe manuscript and for English improvements


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    892 Environ Sci Pollut Res (2012) 19:879892

    Interactions between Zn and bacteria in marine tropical coastal sedimentsAbstractAbstractAbstractAbstractIntroductionMaterial and methodsSampling procedureMicrocosm setupMetal analysisNutrient analysisMicrobial population analysisMathematical analysis

    Results and discussionSediment compositionBacterial community structure and impact of ZnZn adsorption kineticsZn sediment repartitionEffect of zinc spiking on nutrient fluxes