Valérie Feigenbrugel

Stéphane Le Calvé
email: slecalve@illite.u-strasbg.fr

Philippe Mirabel

Florent Louis

Henry's Law Constant measurements for phenol, o-, m-, and p-cresol as a function of temperature

Keywords: Henry's Law Constant, temperature, salinity, phenol, cresols

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Introduction

Phenols constitute a class of organic substances of particular environmental interest. Due to their widespread discharge into the environment and their toxicity to many living organisms, phenol and its methyl derivates (cresols) are presently found on most priority lists of hazardous substances [START_REF] Mackay | Multimedia Environmental Models. The Fugacity Approach[END_REF]. The major emission sources of phenol and cresol isomers into the troposphere are from automobile exhaust, wood burning and industrial sources. These compounds are released to the environment as fugitive emissions and effluents during coal tar and petroleum refining, metal refining and wood pulping as well as during manufacturing and use of many specialized chemicals (resins, disinfectants, insecticides, fungicides, etc.). Phenol and cresols can also be formed in the atmosphere by gas phase photo-oxidation of alkylbenzenes [START_REF] Atkinson | Formation of ring-retaining products from the OH radical-initiated reactions of benzene and toluene[END_REF][START_REF] Bjergbakke | UV spectrum and kinetics of hydroxycyclohexadienyl radicals[END_REF][START_REF] Klotz | Atmospheric oxidation of toluene in a largevolume outdoor photoreactor: In situ Determination of ring-retaining products yields[END_REF][START_REF] Smith | Primary product distribution from the reaction of hydroxyl radicals with toluene at ppb NOx mixing ratios[END_REF][START_REF] Berndt | On the formation of benzene oxide/oxepin in the gas-phase reaction of OH radicals with benzene[END_REF][START_REF] Berndt | Gas-phase reaction of OH radicals with benzene: products and mechanism[END_REF][START_REF] Volkammer | OH-initiated oxidation of benzene: I. Phenol formation under atmospheric conditions[END_REF]. Ambient levels of phenol and cresols have been measured in urban atmosphere (Leuenberger et al., 1985). Results of these studies revealed a mean concentration of 320, 71, and 130 ng m -3 respectively for phenol, o-cresol, and the sum of m-cresol and p-cresol.

Emission data from road traffic indicated that phenol and cresols can reach up to 3% of the overall relative concentrations of the VOCs measured for different locations and engine types [START_REF] Kurtenbach | Trace organic compounds in rain: 4. Identities, concentrations, scavenging mechanisms for phenols in urban air and rain[END_REF][START_REF] Kurtenbach | Trace organic compounds in rain: 4. Identities, concentrations, scavenging mechanisms for phenols in urban air and rain[END_REF][START_REF] Schmitz | Determination of VOC-components in the exhaust of gasoline and diesel passenger cars[END_REF].

Phenol and cresol isomers are polar molecules. They have high solubilities in water and are expected to be efficiently scavenged from the gas phase by rain and fog. Values ranging between 0.2 and 9 µg L -1 were measured for phenol at different locations and the concentration of cresols in the rainwater was estimated to be in the range 0.38 -2 µg L -1 [START_REF] Kawamura | Organic compounds in the rain water of Los Angeles[END_REF]Leuenberger et al., 1985;[START_REF] Levsen | Organic compounds in precipitation[END_REF][START_REF] Richartz | Nitrated phenols in fog[END_REF][START_REF] Grosjean | Atmospheric fate of toxic aromatic compounds[END_REF][START_REF] Tremp | Phenols and nitrophenols as tropospheric pollutants -emissions from automobile exhausts and phase-transfert in the atmosphere[END_REF][START_REF] Allen | Phenol concentrations in air and rain water samples collected near a wood preserving facility[END_REF]Lüttke and Levsen, 1997;Lüttke et al., 1997;[START_REF] Belloli | Determination of toxic nitrophenols in the atmosphere by high-performance liquid chromatography[END_REF][START_REF] Lüttke | Phenols and nitrated phenols in clouds at Mount Brocken[END_REF][START_REF] Morville | Analysis of phenols and nitrophenols in rain and fog collected in an urban area (Strasbourg, France)[END_REF]. These concentration levels make them one of the most abundant organic compounds in rainwater.

An additional atmospheric sink for phenol and cresols is the removal by wet and dry depositions. Phenol, cresols and catechols (o-dihydroxybenzene) were the most common phenolic compounds identified in seawater sampled in coastal California with typical concentrations ranging from 2.5 to 370 ng L -1 [START_REF] Boyd | Identification and quantification of mono-, di-and trihydroxybenzenes (phenols) at trace concentrations in seawater by aqueous acetylation and gas chromatographic-mass spectrometric analysis[END_REF].Biodegradation will generally be the dominant loss mechanism when cresols are released into water with half-lives of several days in estuarine waters. The half-lives are shorter in fresh water and longer in marine water.

An additional source of phenol and cresols in seawater is the produced water issued from the oil production. Methylphenols are an important group of organic contaminates in produced water with concentrations between 0.6 and 10 mg L -1 in the North Sea [START_REF] Meier | Determination of alkylphenols in marine tissues using pentafluorobenzoyl derivatization of phenols and GC/MS (negative-ion chemical ionization)[END_REF].

The partitioning of phenol and cresol isomers between the aqueous liquid phase and the air has to be investigated in order to shed some light in the understanding of the fate, the transport, and the removal of these pollutants in the troposphere. A key physical characteristic to describe the behaviour of a component in the environment [START_REF] Rice | Henry's Law Constants for pesticides measured as a function of temperature and salinity[END_REF] is given by the Henry's law constant (HLC) that represents the partitioning between the gas and aqueous phases.

In order to use HLCs in environmental modeling, temperature and salinity effects have also to be investigated. Although several database of HLCs exist for organic compounds [START_REF] Staudinger | A critical review of Henry's law constants for environmental applications[END_REF][START_REF] Sander | Compilation of Henry's law constants for inorganic and organic species of potential importance in environmental chemistry[END_REF], a detailed inspection of the literature showed a substantial variation in the previously quoted values for phenol and cresols : the data results may differ by as much as a factor of 3.

In this work, direct HLC determinations have been therefore performed with a dynamic equilibrium system, between 278 and 298 K. The effect on HLCs of a 35 g L -1 °/°°°/°° solution of NaCl, used as a surrogate for an ocean seawater, was also studied over the same temperature range.

Experimental Section

Reagents

Aqueous solutions of phenol and cresols were made using Milli-Q Water (18 m cm).

Pure phenol (GC grade,  99.5%), o-cresol (GC grade,  99.5%), m-cresol (GC grade,  98.0%), and p-cresol (GC grade,  98.0%) were obtained from Fluka Chemical Company and used without further purification. Sodium chloride ( 99.8%) was obtained from Riedel-de Haën Chemical Company. The aqueous concentrations used for the experiments were in the range 40-300 mg L -1 for phenol, 80-260 mg L -1 for o-cresol, 50-180 mg L -1 for m-cresol and 60-310 mg L -1 for p-cresol, given that their water solubilities at 298 K are 77500 mg L -1 [START_REF] Frier | Aqueous solutions:Data for Inorganic and Organic Compounds[END_REF], 25900, 22700, and 21520 mg L -1 respectively [START_REF] Yalkowsky | The AQUASOL DATABASE of Aqueous Solubility[END_REF]. To investigate the effects of salinity on HLCs, a water solution of NaCl (35 g L -1 ) was used.

Dynamic equilibrium system

The measurements of HLC were conducted using a dynamic equilibrium system shown in figure 1 and already used in previous work [START_REF] Gautier | Henry's law constants measurements of alachlor and dichlorvos between 283 and 298 K[END_REF][START_REF] Katrib | Uptake measurements of dibasic esters by water droplets and determination of their Henry's law constants[END_REF][START_REF] Feigenbrugel | Henry's law constants measurements of metolachlor and diazinon between 283 and 298 K[END_REF].

Compressed air was passed through a microporous PTFE membrane tube (180 cm  0.8 cm i.d., Sumitomo Corporation) at a highly controlled flow rate (0.1 -0.5 L min -1 ). The tube was immersed in about 0.6 L of the diluted aqueous solution containing the phenolic derivative. Phase equilibrium was achieved, for the dissolved compound, at the gas/water interface along the inner surface of the tube. The contact time between the two phases varied between 7 and 36 s depending on the flow rate. At the exit of the reactor, the gas phase which contains the studied compound at equilibrium concentration was then diluted using clean dry air (0.5 L min -1 ). The compound present in the gas phase was trapped using two cartridges connected in series and filled with a XAD-2 Amberlite resin (Supelco). Before use, this resin was purified by Soxhlet extraction with n-hexane/dichloromethane (50/50) for 24 hours and then dried at room temperature [START_REF] Sanusi | A Multiresidue Method for Determination of Trace Levels of Pesticides in Atmosphere[END_REF][START_REF] Sauret | Analytical method using gas chromatography and ion trap tandem mass spectrometry for the determination of S-triazines and their metabolites in the atmosphere[END_REF].

After sampling, the compound was extracted three times by passing 20 mL, 10 mL and 10 mL of methanol through the cartridge. The mass of methanol which stayed in the cartridge was measured after each extraction in order to know the dead volume in the cartridge. The experiments revealed that for o-cresol more than 99% of the compound was trapped in the first cartridge while for phenol, m-cresol and p-cresol the same percentage ranged respectively between 70 and 100 %, 85 and 100%, and 94 and 100%. 20 minutes to 3 hours were required to concentrate the phenol and cresols in the cartridges. Note that the mass of phenolic derivatives trapped in the cartridges was always negligible compared to their total mass in the aqueous solution so that their aqueous concentration were effectively constant during the course of the experiment.

Analysis

Phenol and cresols were analysed with a GC-MS (CGP 6890N / MSD 5973N, Agilent Technologies) equipped by a 5%-phenylmethylsiloxane capillary column (HP-5MS, 30 m  0.25 mm i.d.  2.5 µm film thickness). For this, 1µL of the solution obtained after extraction was injected in the splitless mode by an automatic sampler (7683, Agilent Technologies).

Helium (UHP certified to > 99.9995% from Air Liquid) was used as carrier gas with a flowrate of 1 mL min -1 and the injector temperature was set at 250°C. For all the solutions, the column temperature was kept at 60°C for 4 min, was then brought to 220°C at a rate of 20°C min -1 , and was held at 220°C for 4 min.

In order to increase the sensitivity, phenol and cresols were analysed by applying single ion monitoring mode of the characteristic fragment ions at m/z 94, 107, and 108 amu. The gas-phase concentrations of the compounds were quantitatively determined from calibration curves made after each measurement. For this, known concentrations of the compounds were prepared in methanol and 1 µL of the obtained solution was then injected into the GC-MS in order to plot concentration curves versus GC-MS peak areas. The calibration curves were linear in the concentration range used in this study i.e. between 100-4000 µg L -1 for phenol, 200-2000 µg L -1 for o-cresol, 200-2000 µg L -1 for m-cresol and 160-3300 µg L -1 for p-cresol and the correlation coefficients were always greater than 0.99.

Results and discussion

Effect of temperature

The Arrhenius plots of the Henry's Law Constants (in M atm -1 ) measured for the phenol and cresols in deionized water for a temperature range 278-298 K are presented in figure 2.

The HLC values are also listed in tables 1 to 4 where the quoted uncertainties vary between 6 and 18 %. These values correspond to the sum of the uncertainties on the calibration curve (ranging from 1 and 5 % for phenol and cresols), on the extraction process (determined to be of the order of 5 %), on the flow rates (less than 2.5 % in our conditions) and on the aqueous concentration of the phenol and cresols (less than 0.5 % for the phenol and cresols due to the relatively high aqueous concentrations in the experiments). Within these experimental errors, HLC values of the phenol and cresols are shown to be independent of their aqueous concentration over the ranges stated in the experimental section (see tables 1 to 4) and of the changes in air flow rate (0.1 -0.5 L min -1 ) as shown in figure 3. This confirms that the equilibrium between solution and vapour is indeed achieved at the exit of reactor.

At 293 K, our recommended values of HLC calculated from the Arrhenius expression using a weighted analysis fit (see text below), are the following (in units of M atm -1 ): HLC = 1005 ± 270 (phenol); HLC = 690 ± 95 (o-cresol); HLC = 1324 ± 172 (m-cresol); HLC = 1742 ± 360 (p-cresol). The quoted errors correspond to 2 level and are determined from the standard deviations on both slope and intercept of Arrhenius expression. These data together with those reported in figure 2 show that the position of the methyl groups on the aromatic ring influences largely the HLC values of the cresols. For instance, the HLC of p-cresol is approximately 3 times higher than that obtained for o-cresol. Moreover, the temperature dependence are very similar for the four studied compounds meaning that their enthalpies of solvation are approximately the same. Based on these observations, it seems that the HLC values of other phenolic derivatives can be predicted at low temperature from a single determination at room temperature.

The following Arrhenius expressions can be derived from our data using a weighted analysis fit: HLC(phenol) = (4.1 ± 0.6)  10 -9 exp((7684 ± 874)/T), HLC(o-cresol) = (1.5 ± 0.1)  10 -10 exp((8544 ± 512)/T), HLC(m-cresol) = (5.5 ± 0.4)  10 -11 exp((9028 ± 508)/T) and HLC(p-cresol) = (3.3 ± 0.4)  10 -11 exp((9258 ± 818)/T). The quoted errors are given at the 2 level from the least-squares analysis. These Arrhenius expressions permit to highlight the strong temperature dependence of HLC values for phenol and cresols. For instance, at 283 K, HLC values are between 2.5 and 3 times higher that those measured at 293 K. This is in quite good agreement with the commonly cited "rule of thumb" [START_REF] Staudinger | A critical compilation of Henry's law constant temperature dependence relations for organic compounds in dilute aqueous solutions[END_REF] that doubles HLC values per 10°C decrease in temperature.

In Tables 5678, our values of HLC are compared with those given in the literature. Tables 5-8 also provide previously reported temperature dependence of HLC whenever available.

One note that few measurements, as a function of temperature, are available for m-and pcresols and therefore the comparison is rather limited in these cases.

Concerning phenol, our value at 293 K is in quite good agreement with that determined by [START_REF] Tremp | Phenols and nitrophenols as tropospheric pollutants -emissions from automobile exhausts and phase-transfert in the atmosphere[END_REF] using a Stripping Column Technique (see table 5) and with that of [START_REF] Sheikheldin | Determination of Henry's law constants of phenols by pervaporation-.ow injection analysis[END_REF] Our experimental data are also quite consistent with the average value of 9206 M atm -1 derived at 278 K from simultaneous measurements of the phenol concentration in gas and aqueous phases (Lüttke and Levsen, 1997). However, our values are between 2 and 5 times lower than the previous ones [START_REF] Parsons | Effect of 4-substitution on the thermodynamics of hydration of phenol and the phenoxide anion[END_REF][START_REF] Abd-El-Bary | Henry's constants for phenol over its diluted aqueous solution[END_REF][START_REF] Sheikheldin | Determination of Henry's law constants of phenols by pervaporation-.ow injection analysis[END_REF][START_REF] Harrison | Experimentally determined Henry's Law coefficients of phenol, 2-methylphenol and 2-nitrophenol in the temperature range 281-302[END_REF] as shown in figure 4. On the contrary, the value reported by [START_REF] Janini | Determination of activity coefficients of oxygenated hydrocarbons by liquid-liquid chromatography[END_REF] at 296 K is approximately 4 times lower than our own determination.

For o-cresol, our experimental determination of HLC at 293 K agrees well, within experimental errors, with the values reported by [START_REF] Tremp | Phenols and nitrophenols as tropospheric pollutants -emissions from automobile exhausts and phase-transfert in the atmosphere[END_REF] and [START_REF] Sheikheldin | Determination of Henry's law constants of phenols by pervaporation-.ow injection analysis[END_REF], who determined HLC by using a stripping column technique and a single equilibrium static technique, respectively (see table 6). The agreement is also reasonable between our measurement at 298 K and the values reported by [START_REF] Parsons | The thermodynamics of hydration of phenols[END_REF] and [START_REF] Altschuh | Henry's law constants for a diverse set of organic chemicals: Experimental determination and comparison of estimation methods[END_REF] while our determination is 2.5 times lower than that determined by [START_REF] Harrison | Experimentally determined Henry's Law coefficients of phenol, 2-methylphenol and 2-nitrophenol in the temperature range 281-302[END_REF] and 2 times higher than that reported by [START_REF] Janini | Determination of activity coefficients of oxygenated hydrocarbons by liquid-liquid chromatography[END_REF].

For m-cresol, our determination at 293 K is in very good agreement with the previous one reported recently by [START_REF] Sheikheldin | Determination of Henry's law constants of phenols by pervaporation-.ow injection analysis[END_REF] as shown in table 7. Our value at 298 K is also consistent with that determined by Altschuch et al. (1999). However, our value at 293 K is about 2 times higher than the previously reported data by Janini and Quaddora [START_REF] Janini | Determination of activity coefficients of oxygenated hydrocarbons by liquid-liquid chromatography[END_REF] who estimated HLC from experimentally determined activity coefficients and reported vapor pressure data.

For p-cresol, our experimental determination of HLC at 293 K in deionized water is consistent with the lower limit estimated by Altschuh et al. and the values obtained by [START_REF] Parsons | The thermodynamics of hydration of phenols[END_REF] and by [START_REF] Tremp | Phenols and nitrophenols as tropospheric pollutants -emissions from automobile exhausts and phase-transfert in the atmosphere[END_REF] even if these latter are about 30% lower (see table 8). Our value of HLC obtained at 293 K is about more than 3 times higher than the previous one reported by [START_REF] Janini | Determination of activity coefficients of oxygenated hydrocarbons by liquid-liquid chromatography[END_REF].

For phenol and o-cresol, we report respectively the fifth and the fourth temperature dependence of HLC. To our knowledge, we report the second measurement of HLC as a function of temperature for m-and p-cresol. The previous temperature dependence recommended by [START_REF] Janini | Determination of activity coefficients of oxygenated hydrocarbons by liquid-liquid chromatography[END_REF] is probably suspicious because their determination at 293 K are much lower than all the other reported values for the four studied compounds. For phenol and o-cresol, our temperature dependence d(lnHLC)/d(1/T) agrees reasonably with the previous determinations [START_REF] Parsons | Effect of 4-substitution on the thermodynamics of hydration of phenol and the phenoxide anion[END_REF][START_REF] Abd-El-Bary | Henry's constants for phenol over its diluted aqueous solution[END_REF][START_REF] Harrison | Experimentally determined Henry's Law coefficients of phenol, 2-methylphenol and 2-nitrophenol in the temperature range 281-302[END_REF].

Effect of salinity

The Arrhenius plots of the Henry's Law Constants, measured for the phenol and cresols in aqueous solutions of NaCl (35 g L -1 ) simulating seawater for a temperature range 278-298 K, are presented in the figures 4-7. The following Arrhenius expressions can be derived from our data using a weighted analysis fit: HLCS = (1.2 ± 0.4)  10 -10 exp((8619 ± 2240)/T), HLCS = (3.4 ± 0.6)  10 -11 exp((8905 ± 1144)/T), HLCS = (1.3 ± 0.4)  10 -10 exp((8710 ± 1854)/T) and HLCS = (6.8 ± 1.8)  10 -12 exp((9651 ± 1920)/T) for phenol, o-cresol, m-cresol and p-cresol, respectively. All of the HLC values for the salt solution were 10 to 30 % lower than their respective values in deionized water (Tables 1234), depending on the compound and the temperature. For instance, the determined HLC values at 293 K in salt solution obtained from Arrhenius expression given above, can be compared with those measured in pure water (units of M atm -1 ). This phenomenon is related to the salting-out effect which impacts on the water solubility (reduction in solubility by a factor of 1.25 to 1.40 for the studied compounds).

The effect of salt on the solubility of gases in seawater has been studied theoretically for a variety of small gas molecules [START_REF] Danckwerts | Gas-liquid reactions[END_REF]. The presence of a salt in an aqueous solution changes the water structure in the vicinity of the solute. An empirical relation for the effect of ionic strength on a compound's HLC in an aqueous solution was given by [START_REF] Setschenow | Über die Konstitution der Salzlösungen auf Grund ihres Verhaltens zu Kohlensäure[END_REF]:

log (HLC / HLCS) = h (NaCl) (1)
where HLC and HLCS are the Henry's law coefficient of the compound in pure water and in the salt solution, respectively, (NaCl) is the molarity of the NaCl solution and h is the Setschenow salting-out coefficient (L mol -1 ).

Our results of HLC and HLCS at 293 K (see tables 5-8) were used to estimate the following Setschenow coefficients: phenol, 0.25; o-cresol, 0.19; m-cresol, 0.16; p-cresol, 0.18. Values of the salting-out coefficients are similar to those of other phenolic derivatives [START_REF] Xie | A Review of the Effects of Salts on the Solubility of Organic Compounds in Seawater[END_REF]. Our determinations are also close to those obtained from experiments for phenol by [START_REF] Morrison | The salting-out effect[END_REF] (0.22 for (NaCl) = 1 M at 298 K) and m-cresol by [START_REF] Carter | The salting-out effect: Influence of electrolytes on the solubility of m-cresol in water[END_REF]) (0.17 for (NaCl) = 0.5 M at 298 K).

These data on salting-out are valuable for assessing the solubility of organic substances in seawater from corresponding data for freshwater. The magnitude of the salting-out effect is large enough to have a significant influence on mass transfer rate of phenolic compounds in natural processes. For example, at 293 K and for an ionic strength of 0.6 M, which is representative of seawater, the rate of absorption of the gases into the ocean from the troposphere would be slower than in freshwater of the same temperature.

Atmospheric implications

Once released in the atmosphere, Volatile Organic Compounds (VOCs) can react in the gas phase or can be taken up by water droplets where they can also be oxidized. For most of them, OH radical is the most efficient oxidizing species in both the gas phase [START_REF] Atkinson | Atmospheric chemistry of VOCs and NOx[END_REF] and the aqueous phase [START_REF] Ervens | [END_REF]. To evaluate the impact of a cloud on the atmospheric chemistry of phenol and cresols, we compare in this section their atmospheric lifetimes under clear sky (τgas), and cloudy conditions (τmultiphase).

The lifetime of A in the gas phase (τA,gas) taking only into account the reaction of A with OH radicals, is defined as follow:

g g OH, g g gas A, ] [OH k 1 /dt d[A] [A] τ = - = (3)
where kOH,g is the rate constant of the gas phase reaction of OH radicals with A, [OH]g is the mean concentration of OH radicals and [A]g is the concentration of A in the troposphere.

If we consider now the cloudy atmosphere as a multiphase reactor where both the gas and aqueous phases coexist, then the multiphase lifetime of A can be defined as [START_REF] Monod | Kinetics of OHinitiated oxidation of some oxygenated organic compounds in the aqueous phase under tropospheric conditions -1 -New results[END_REF]:

/dt d[A] - /dt d[A] [A] [A] aq g aq g multiphase A, τ - + = (4)
where [A]aq is the concentrations of A in the aqueous phase. At equilibrium, the fractions of A in the gas (fA,g) and aqueous phases (fA,aq) are the following: (5)

T L R HLC 1 T L R HLC [A] [A] [A]
T L R HLC 1 1 [A] [A] [A] wc aq g g , + = + = g A f (6)
where R the ideal gas constant (in L atm mol -1 K -1 ), Lwc the dimensionless liquid water content of the cloud (typically 4.210 -7 ).

Mass accommodation can limit the rate of gas uptake. The accommodation time-scale accom can be estimated from the mass accommodation coefficient  as:

α A c 4 1 1 τ c accom   = ( 7
)
where <c> is the mean molecular velocity of the trace gas.

The mass accommodation coefficients  have been measured for the cresol isomers by Louis and coworkers [START_REF] Leyssens | Temperature Dependence of the Mass Accommodation Coefficients of 2-Nitrophenol, 2-Methylphenol, 3-Methylphenol, and 4-Methylphenol on Aqueous Surfaces[END_REF] and their values at 293K are respectively equal to 6.6  10 -4 for o-cresol, 1.1  10 -3 for m-cresol, and 2.9  10 -3 for p-cresol. In addition, a value of 8.3  10 -3 for phenol at 293 K has already been reported by [START_REF] Müller | Mass Accommodation Coefficients of Phenol, 2-Nitrophenol, and 3-Methylphenol over the Temperature Range 278-298 K[END_REF]. From the values of  the calculated accommodation time scale varies between 0.6 and 8.4 min. In regard to the tropospheric lifetimes of phenol or cresols in the gas phase (see table 9), mass accommodation is indeed a fast process. Therefore, the uptake of these compounds by atmospheric droplets will not be limited by mass transport.

Assuming that equilibrium is effectively rapidly reached, eq. ( 3) can be written as follows:

) ] [OH (k T L R HLC ) ] [OH (k T L R HLC 1 τ aq OH,aq wc g OH,g wc multiphase A,  + + = (8)
where kOH,aq is the rate constant of OH reaction with A in aqueous phase, [OH]aq is the mean concentration of OH radicals in the tropospheric aqueous phase.

The rate constants of OH reaction with phenol or cresols in both aqueous and gas phases have already been reported [START_REF] Savel'eva | Reactions of substituted phenols with hydroxyl radicals and their dissociated form O[END_REF][START_REF] Feitelson | [END_REF][START_REF] Field | A Pulse Radiolysis Investigation of the Reactions of BrO,. with Fe(CN),4-, Mn( I I ), Phenoxide Ion, and Phenol[END_REF][START_REF] Semadeni | The Temperature Dependence of OH Radical Reactions with Aromatic Compounds under Simulated Tropospheric Conditions[END_REF], and are shown in table 9 together with the Henry's law constants determined in this work. By analogy with other cresols, the rate coefficient of OH reaction with m-cresol was estimated to 1.210 10 M -1 s -1 . These data have then been used to estimate the atmospheric lifetimes of these organic compounds under clear sky (τgas), and cloudy conditions (τmultiphase), by using typical concentrations of OH radicals in the gas and in aqueous phases: [OH]g=110 6 cm -3 and [OH]aq=110 -13 M [START_REF] Monod | Kinetics of OHinitiated oxidation of some oxygenated organic compounds in the aqueous phase under tropospheric conditions -1 -New results[END_REF]. We assume here that the OH concentration in the gas phase is the same under clear sky and cloudy conditions although it is reduced as soon as clouds are formed, and the magnitude of the decrease strongly depends on the pH value of the droplet [START_REF] Monod | Impact of clouds on the tropospheric ozone budget: Direct effect of multiphase photochemistry of soluble organic compounds[END_REF].

At 298 K, the calculated aqueous fraction are the following: phenol, 6.710 -3 ; o-cresol, 4.310 -3 ; m-cresol, 8.110 -3 ; p-cresol, 10.610 -3 . At 298 K which is the temperature where most of experimental rate constants were measured, the atmospheric lifetimes of phenol and cresols are slightly reduced in the cloud as shown in table 9. However, the average temperature of tropospheric clouds is 283 K and therefore, the calculated aqueous fraction of phenol rises from 6.710 -3 to 2.810 -2 when the temperature decreases from 298 K to 283 K, temperature range where the HLC value of phenol increases from 658 to 2815 M atm -1 .

Assuming that the rate constants of OH reactions with phenol or cresols in the gas and aqueous phases do not vary a lot between 283 and 298 K as it was already observed for many oxygenated VOCs, we can then estimate the multiphase lifetimes of these compounds at 283 K. These calculated lifetimes at 283 K (in units of days) are significantly lower than those at 298 K (in parentheses): phenol, 0.26 (0.38); o-cresol, 0.17 (0.22); m-cresol, 0.13 (0.19); pcresol, 0.11 (0.18). Considering all the uncertainties associated with the above calculations, one can conclude that the oxidations of phenol and cresols in the gas and aqueous phases are of the same order of magnitude and can really compete. Mechanisms of OH-oxidation in aqueous phase are therefore needed to identify the products that can be deposited to ground by wet deposition and consequently to better estimate the environmental impact of phenol and cresols.

Figure Captions

Figure 1 : Scheme of the dynamic equilibrium system used to determine Henry's law constants as a function of the temperature. and with an aqueous concentration of o-cresol of  150 mg L -1 ( 1.410 -3 M). Experiments carried out with an air flow rate of: a 0.1 L min -1 ; b 0.3 L min -1 ; c 0.4 L min -1 ; d 0.5 L min -1 .

Experiments performed with an aqueous o-cresol concentration of: e  100 mg L -1 ( 9.310 -4 M); f  200 mg L -1 ( 1.910 -3 M). ); g  250 mg L -1 ( 2.310 -3 M). h Salt water experiments (35 g L -1 °/°°). [START_REF] Janini | Determination of activity coefficients of oxygenated hydrocarbons by liquid-liquid chromatography[END_REF][START_REF] Parsons | The thermodynamics of hydration of phenols[END_REF][START_REF] Tremp | Phenols and nitrophenols as tropospheric pollutants -emissions from automobile exhausts and phase-transfert in the atmosphere[END_REF] Figure 7

(

  in parentheses): phenol, HLC = 715 ± 430 (1005 ± 270); o-cresol, HLC = 532 ± 157 (692 ± 95); m-cresol, HLC = 1057 ± 527 (1324 ± 172); p-cresol, HLC = 1369 ± 630 (1742 ± 360)

Figure 2 :

 2 Figure 2 : Plot of ln HLC versus 1000/T in deionized water for the four compounds studied: phenol, o-cresol, m-cresol, p-cresol, in the temperature range of 278-298 K. Our errors on HLC varied typically between 6 and 18 % and are determined as mentioned in the text.

Figure 3 :

 3 Figure 3 : Plot of HLC versus the flow rate ranging between 0.1 and 0.5 L min -1 for phenol, o-cresol, m-cresol, p-cresol at 293 K. Our errors on HLC varied typically between 6 and 11 % and are determined as mentioned in the text.

Figure 4 :

 4 Figure 4 : Plot of ln HLC versus 1000/T for phenol in the temperature range of 278-298 K in deionized and salt water. Our determinations are compared to the experimental data of the literature. Our errors on HLC varied typically between 7 and 9 % and are determined as mentioned in the text. Straigth lines represent Arrhenius plots derived by Harrison et al. (long dash line), by Abd-El-Bary et al. (short dash line), or in this work for deionized water (solid line) and a 35 g L -1 °/°° solution of NaCl (dash-dot line).

Figure 5 :

 5 Figure 5 : Plot of ln HLC versus 1000/T for o-cresol in the temperature range of 278-298 K in deionized and salt water. Our determinations are compared to the experimental data of the literature. Our errors on HLC varied typically between 6 and 12 % and are determined as mentioned in the text. Straigth lines represent Arrhenius plots derived by Harrison et al. (long dash line), or in this work for deionized water (solid line) and a 35 g L -1 °/°° solution of NaCl (dash-dot line).

Figure 6 :

 6 Figure 6 : Plot of ln HLC versus 1000/T for m-cresol in the temperature range of 278-298 K in deionized and salt water. Our determinations are compared to the experimental data of the literature. Our errors on HLC varied typically between 8 and 11 % and are determined as mentioned in the text. Solid and dash-dot lines represent Arrhenius plots derived in this work for deionized water and a 35 g L -1 °/°° solution of NaCl, respectively.

Figure 7 :

 7 Figure 7 : Plot of ln HLC versus 1000/T for p-cresol in the temperature range of 278-298 K in deionized and salt water. Our determinations are compared to the experimental data of the

  Figure1

Table 1 :

 1 Henry's Law Constants of phenol as a function of the temperature determined by using the dynamic equilibrium system.

	T (K)	HLC (M atm -1 )
	278.15 f	4566 ± 355
	278.15 h	3390 ± 252

Table 2 :

 2 Henry's Law Constants of o-cresol as a function of the temperature determined by using the dynamic equilibrium system.

	T (K)	HLC (M atm -1 )
	278.25	2912 ± 280
	278.25 h	2858 ±240
	278.35	3548 ± 290
	283.15 g. h	1434 ± 107
	283.15 e	1824 ± 179
	283.25 e	1792 ± 212
	288.10	1204 ± 103
	288.15 f. h	891 ± 67
	288.20	1185 ± 100
	293.10 e	632 ± 58
	293.15 c. e	623 ± 48
	293.15 d. g	652 ±42
	293.15 a	739 ± 76
	293.15 h	538 ± 45
	293.25 b	659 ± 49
	298.15 f	430 ± 33
	298.25 e	429 ± 40
	By default, the experiments were carried out in deionized water with a flow rate of 0.2 L min -1

Table 3 :

 3 Henry's Law Constants of m-cresol as a function of the temperature determined by using the dynamic equilibrium system.

	T (K)	HLC (M atm -1 )
	278.15	6864 ± 568
	278.25 h	5268 ± 519

Table 4 :

 4 Henry's Law Constants of p-cresol as a function of the temperature determined by using the dynamic equilibrium system.

	T (K)	HLC (M atm -1 )
	278.25 e	11090 ± 1998
	278.25 h. j	8829 ± 825

Table 5 :

 5 Comparison of experimental Henry's law constants for phenol with previous work.

	T (K)	H (M atm -1 )	d ln H / d(1/T) (K -1 )	T range (K)	Technique	Reference
	278	9206 g	-	-	FE a	(Lüttke and Levsen, 1997)
	291.5	2709	6375	271-363	DES b	(Abd-El-Bary et al., 1986)
	293	227	3600	298-308	  c	(Janini and Quaddora, 1986)
	293	1559	-	-	SCT d	(Tremp et al., 1993)
	293	1858 ± 491	-	-	SEST e	(Sheikheldin et al., 2001)
	293	3870	5850	284-302	SCT d	(Harrison et al., 2002)
	298	>421 h	-	-	DES b	(Altschuh et al., 1999)
	298	2900	6800	230-273	Pv f	(Parsons et al., 1971)
	278 4138 ± 1113	7684 ± 874	278-298	DES b	This work, deionized water
	291.5 1150 ± 309
	293	1005 ± 270
	298	647 ± 174
	293	715 ± 430	8619 ± 2240	278-293	DES b	This work, salt water

a HLC derived from field experiments; b DES: Dynamic Equilibrium System; c   : Infinite dilution activity in water; d SCT: Stripping Column Technique; e SEST: Single Equilibrium Static Technique; f Pv: Calculated value from the ratio between the vapour pressure, experimentaly measured and the solubility; g average value; h Not used in validation analysis by the author.

Table 6 :

 6 Comparison of experimental Henry's law constants for o-cresol with previous work.

	T (K)	H (M atm -1 )	d ln H / d(1/T) (K -1 )	T range (K)	Technique	Reference
	293	338	4600	298-308	  a	(Janini and Quaddora, 1986)
	293	500 ± 116	-	-	SEST b	(Sheikheldin et al., 2001)
	293	724	-	-	SCT c	(Tremp et al., 1993)
	293	1560	6680	281-302	SCT d	(Harrison et al., 2002)
	298	830	-	-	Sol. a	(Parsons et al., 1972)
	298	638	-	-	DES e	(Altschuh et al., 1999)
	293	692 ± 95	8544 ± 512	278-298	DES e	This work, deionized water
	298	424 ± 58
	293	532 ± 157	8905 ± 1144	278-293	DES e	This work, salt water

a   : Infinite dilution activity in water; b SEST: Single Equilibrium Static Technique; c SCT: Stripping Column Technique; d Sol.: Calculated value from the ratio between the vapour pressure and the solubility, experimentaly measured; e DES: Dynamic Equilibrium System.

Table 7 :

 7 Comparison of experimental Henry's law constants for m-cresol with previous work.

	T (K)	H (M atm -1 )	d ln H / d(1/T) (K -1 )	T range (K)	Technique	Reference
	293	630	7700	298-308	  a	(Janini and Quaddora, 1986)
	293 1239  199	-	-	SEST b	(Sheikheldin et al., 2001)
	298	1168	-	-	DES c	(Altschuh et al., 1999)
	293 1324 ± 172	9028 ± 508	278-298	DES c	This work, deionized water
	298	798 ± 102
	293 1057 ± 527	8710 ± 1854	278-293	DES c	This work, salt water

a   : Infinite dilution activity in water; b SEST: Single Equilibrium Static Technique; c DES: Dynamic Equilibrium System.

Table 8 :

 8 Comparison of experimental Henry's law constants for p-cresol with previous work.

	T (K)	H (M atm -1 )	d ln H / d(1/T) (K -1 )	T range (K)	Technique	Reference
	293	530	4600	298-308	  a	(Janini and Quaddora, 1986)
	293	1351	-	-	SCT b	(Tremp et al., 1993)
	298	1300	-	-	Sol. c	(Parsons et al., 1972)
	298	> 292 e	-	-	DES d	(Altschuh et al., 1999)
	293 1742 ± 360	9258 ± 818	278-298	DES d	This work, deionized water
	298 1025 ± 212
	293 1369 ± 630 9651 ± 1920	278-293	DES d	This work, salt water

a   : Infinite dilution activity in water; b SCT: Stripping Column Technique; c Sol.: Calculated value from the ratio between the vapour pressure and the solubility, experimentaly measured; d DES: Dynamic Equilibrium System; e Not used in validation analysis by the author.

Table 9 :

 9 Rate constants of OH reaction with phenol and cresols in gas and aqueous phases and the corresponding Henry's law constants. The atmospheric lifetimes were calculated in cm 3 molecule -1 s -1 ; b values reported by[START_REF] Semadeni | The Temperature Dependence of OH Radical Reactions with Aromatic Compounds under Simulated Tropospheric Conditions[END_REF]; c in M -1 s -1 ; d Field et al. (1982); e Savel'eva et al. (1972); f Feitelson and Hayon (1973); g our experimental values.

	Compound	k	K g OH, 298	a,b	k	298 OH,	K aq	c	H 298 K (M atm -1 )	 gas (days)	 multiphase (days)
	Phenol	2.6  10 -11	6.710 9 d	627 g	0.45	0.38
	o-cresol	4.9  10 -11	1.110 10 e	433 g	0.24	0.22
	m-cresol	5.2  10 -11		-			791 g	0.22	0.19
	p-cresol	5.2  10 -11	1.210 10 f	1101 g	0.23	0.18

a

Acknowledgment

Financial support for this work has been provided by the CNRS through the "Programme National de Chimie Atmosphérique (PNCA)". This work was also sponsored in part by the French Agency for Environment and Energy Management (ADEME) and the region of Alsace. This is the EOST contribution n° 2003.101-UMR7517.


Edwige Gautreau
email: edwige.gautreau@univ-lyon1.fr

Laurence Volatier
email: laurence.volatier@entpe.fr

Géraldine Nogaro
email: geraldine.nogaro@edf.fr

Pierre Rossignol

Laurent Simon

Michel Lafont

Influences of bioturbation and water column oxygenation on nutrient recycling from reservoir sediments Authors

Keywords: water-sediment interface, aerobic/anaerobic conditions, chironomid larvae, tubificid worms, fauna activity, nitrogen, phosphorus

Sediments act as sinks of nutrients in aquatic ecosystems but may also act as nutrient sources, leading to eutrophication of lakes and reservoirs during the warm season of the year. In this context, internal recycling of nutrients from sediments can be largely modulated by abiotic (e.g., redox conditions) and biotic (e.g., activity of benthic organisms) factors occurring at the water-sediment interface. The present study aimed at quantifying the effects of these factors (water column oxygenation and bioturbation by two bioturbating species -the tubificid worm Tubifex tubifex and insect larvae of Chironomus plumosus) on benthic fluxes of nitrogen (N), phosphorus (P) and silicon (Si) from reservoir water-sediment interface. An experimental approach based on the reconstitution of the water-sediment interface in mesocosms has been developed in the laboratory to test three fauna conditions (no fauna, presence of worms, and presence of chironomids) and three conditions of water column oxygenation (constant aerobic conditions, fluctuations of oxygen concentrations and constant anaerobic conditions). The larvae of chironomids significantly increased by 3.7-fold and by 17-fold the concentrations of N (NH4 + + NO3 -) and PO4 3- released from sediments, respectively. In comparison, tubificid worms had lower influences on these released nutrients (x2 for N and x3 for PO4 3-in comparison with the control treatment without fauna). These contrasted effects of chironomids and tubificid worms on nutrient fluxes were related to their different bioturbation activities in sediments. Chironomid larvae increased water-sediment exchanges by actively ventilating their U-shaped burrows whereas tubificid worms produced deep galleries but they did not ventilate them to the same extent as chironomid burrows. Anaerobic conditions increased by 56-fold N fluxes and by 102-fold PO4 3-fluxes in comparison with the aerobic treatment. While anaerobic conditions could produce higher stimulation of N fluxes than bioturbation process, oxic-anoxic shift was also a stronger regulator of P fluxes than benthic fauna. Overall, bioturbating fauna and occurrence of anoxic conditions at the water-sediment interface should not be neglected in model assessing the role of sediments on nutrient dynamics in lakes and reservoirs during the season favorable to algal blooms.

Introduction

The increasing anthropogenic discharge of nutrients is the primary cause of eutrophication in aquatic ecosystems. This phenomenon is predominantly observed in aquatic systems such as lakes and reservoirs [START_REF] Vitousek | Human alteration of the global nitrogen cycle: sources and consequences[END_REF][START_REF] Smith | Eutrophication of freshwater and coastal marine ecosystems a global problem[END_REF][START_REF] Galloway | Nitrogen cycles: past, present, and future[END_REF]. Indeed, reservoirs participate to more than 50% of the sediment retention in drainage basins around the world (4-5 gigatons/year, [START_REF] Vörösmarty | Anthropogenic sediment retention: major global impact from registered river impoundments[END_REF], leading to a massive accumulation of sediment-associated nutrients (nitrogen N and phosphorus P).

During the warm season of the year, the eutrophication of aquatic ecosystems leads to algal blooms, especially cyanobacteria blooms resulting from oxygen depletion in ecosystems [START_REF] Schindler | Eutrophication and recovery in experimental lakes: implications for lake management[END_REF][START_REF] Findlay | Phytoplankton community responses to nutrient addition in Lake 226, Experimental Lakes Area, northwestern Ontario[END_REF], which can lead to effects on fishes and macroinvertebrates (i.e. reduction of populations, [START_REF] Breitburg | Nutrient enrichment and fisheries exploitation: interactive effects on estuarine living resources and their management[END_REF].

For example, [START_REF] Foley | Long-term changes in oxygen depletion in a small temperate lake: effects of climate change and eutrophication: Oxygen depletion in a small lake[END_REF] reported decreasing dissolved oxygen (DO) concentration with eutrophication in a lake monitored during more than 40 years. Consequently, eutrophication significantly impacts the human activities linked to reservoirs (e.g. drinking purposes, fishing) and eutrophic aquatic ecosystems have difficulties to meet EU Water Framework Directive (WFD 2000/60/EC ;European Council, 2000) requirements, which have to reach the good ecological status by 2027 for all surface waters.

In reservoirs, excessive nutrient loadings leading to the eutrophication of the water column have 2 origins: external (tributaries, atmospheric deposition, surface run-off, groundwater [START_REF] Meinikmann | Phosphorus in groundwater discharge -A potential source for lake eutrophication[END_REF]) and internal sources from the sediment compartment. In multiple cases, a reduction of external inputs did not efficiently reduce eutrophication because of internal sources associated with the massive amounts of nutrients accumulated in the sediments, as shown for phosphorus (P) by [START_REF] Søndergaard | Role of sediment and internal loading of phosphorus in shallow lakes[END_REF] or for nitrogen (N) by [START_REF] Jeppesen | Lake responses to reduced nutrient loading -an analysis of contemporary long-term data from 35 case studies[END_REF].

For example, the modelling approach developed by [START_REF] Wu | Internal cycling, not external loading, decides the nutrient limitation in eutrophic lake: A dynamic model with temporal Bayesian hierarchical inference[END_REF] on the eutrophic lake Dianchi (China) showed that internal loading contributed to 77% for total P and 72% for total N of the total inputs to the water column compared to external loading. Moreover, the contribution of internal sources of nutrients from sediments to water column eutrophication was the highest during the warm season of the year. Indeed, the release rate of P from lake sediments to the water column can be more than 12-fold higher in late spring and summer than during other seasons [START_REF] Penn | Seasonality in phosphorus release rates from the sediments of a hypereutrophic lake under a matrix of pH and redox conditions[END_REF][START_REF] Qin | Estimating internal P loading in a deep water reservoir of northern China using three different methods[END_REF].

For internal inputs, molecular diffusion is classically considered as the main mechanism influencing nutrient fluxes from the sediments to the water column [START_REF] Berner | Early diagenesis: a theoretical approach[END_REF][START_REF] Anschutz | Effects of tubificid worm bioturbation on freshwater sediment biogeochemistry[END_REF]. Nevertheless, these fluxes are also known to be influenced by chemical (temperature, redox conditions, pH), physical (hydrodynamics like water currents or wind) and biological (bioturbation) factors [START_REF] Kristensen | Organic matter diagenesis at the oxic/anoxic interface in coastal marine sediments, with emphasis on the role of burrowing animals[END_REF][START_REF] Lavery | The use of Fick's First Law for predicting porewater nutrient fluxes under diffusive conditions[END_REF][START_REF] Ni | Historical accumulation and environmental risk of nitrogen and phosphorus in sediments of Erhai Lake, Southwest China[END_REF].

For instance, the solubility of phosphate (PO4 3-) ions is determined by oxygen availability due to their affinity with iron oxides. The PO4 3-ion can be adsorbed on the precipitated Fe(OH)3-PO4 complexes in presence of oxygen [START_REF] Mortimer | The exchange of dissolved substances between mud and water in lakes[END_REF][START_REF] Boström | Exchange of phosphorus across the sedimentwater interface[END_REF]Wetzel, 2001a). Phosphates can also precipitate with calcite (CaCO3) in alkaline condition (as seen by [START_REF] House | Total phosphorus content of river sediments in relationship to calcium, iron and organic matter concentrations[END_REF]. Bioturbation by benthic animals (i.e. sediment reworking and irrigation of biogenic structures such as tubes and burrows) also influences N and P dynamics at the water sediment interface. By modifying redox potentials and increasing water exchanges at the water-sediment interface, burrowing animals can stimulate the release of N (ammonium and/or nitrate) from interstitial water to the water column (e.g. [START_REF] Krantzberg | The influence of bioturbation on physical, chemical and biological parameters in aquatic environments: a review[END_REF]. For example, [START_REF] Pelegri | Nitrogen cycling in lake sediments bioturbated by Chironomus plumosus larvae, under different degrees of oxygenation[END_REF] reported a 5.8-fold stimulation of ammonium (NH4 + ) release rates from lake sediments due to the bioturbation activities of Chironomus plumosus (2,000 larvae.m -²) in microcosms. Using a comparable experimental approach, [START_REF] Mermillod-Blondin | Do tubificid worms influence the fate of organic matter and pollutants in stormwater sediments?[END_REF] showed that tubificid worms (20,000 individuals.m -²) increased the releases of NH4 + , and PO4 3-from sediments by 2-and 4-fold, respectively. The influence of benthic fauna on nutrient fluxes at the water-sediment interface depends on their functional traits (building of biogenic structures, ventilation rate, depth of biogenic structures, sediment reworking rate) [START_REF] Kristensen | What is bioturbation? The need for a precise definition for fauna in aquatic sciences[END_REF]. For example, [START_REF] Nogaro | Alum application, invertebrate bioturbation, and sediment characteristics interact to affect phosphorus exchange in eutrophic ecosystems[END_REF] showed that chironomid larvae that build and ventilate U-shaped tubes stimulated P release from the sediment to the overlying water, whereas tubificids that produce gallery and egest fecal pellets at the water sediment interface increased P retention in sediment from eutrophic wetlands. Although the presence of Chironomus plumosus can contribute to release of nutrients into water column, they also bring oxygen into sediment and create oxic conditions nearby their burrow walls. These oxic conditions lead to trap phosphates in Fe(OH)3-PO4 complexes inside sediments [START_REF] Hupfer | Chironomid larvae enhance phosphorus burial in lake sediments: Insights from long-term and short-term experiments[END_REF].

Based on these findings, our study aimed to quantify the relative contribution of bioturbation and oxygen concentration conditions on nutrient fluxes at the water-sediment interface of a eutrophic reservoir during the warm season of the year when low oxygen conditions were expected to occur at the water-sediment interface. Laboratory experimentations in mesocosms mimicking the water-sediment interface of the studied reservoir were developed for evaluating the influences of bioturbating fauna (chironomid larvae, tubificid worms) and oxygen conditions in the water column (aerobic, anaerobic and fluctuating conditions) on nutrient (N-NH4 + , N-NO3 -, P-PO4 3-and SiO2) released from sediments. For bioturbating fauna, oxygen uptake was also measured in the experiment as animals can stimulate nutrient released from sediment by stimulating organic matter mineralization in sediments [START_REF] Baranov | Effects of bioirrigation of non-biting midges (Diptera: Chironomidae) on lake sediment respiration[END_REF][START_REF] Murniati | Effects of bioirrigation on the spatial and temporal dynamics of oxygen above the sediment-water interface[END_REF][START_REF] Saaltink | Respiration and aeration by bioturbating Tubificidae alter biogeochemical processes in aquatic sediment[END_REF]. We expected that bioturbating fauna would significantly increase the release of nutrients from sediments [START_REF] Mermillod-Blondin | Influence of bioturbation by three benthic infaunal species on microbial communities and biogeochemical processes in marine sediment[END_REF]. Moreover, due to higher ventilation of biogenic structures, the larvae of Chironomus plumosus would generate higher release rates of nutrients and oxygen uptake than tubificid worms [START_REF] Nogaro | Alum application, invertebrate bioturbation, and sediment characteristics interact to affect phosphorus exchange in eutrophic ecosystems[END_REF]. Concerning DO conditions, anaerobic conditions would enhance the release of nutrients, especially N-NH4 + and P-PO4 3-, compared with aerobic conditions. For the oxygen fluctuation treatment, we reproduced the daily variations of DO concentrations in the water column of eutrophic reservoirs (e.g. [START_REF] Balangoda | Effect of diurnal variation of dissolved oxygen in a eutrophic polymictic reservoir[END_REF]Wetzel, 2001b) by using the natural oxygen uptake of the sediment column without oxygen supply. Thus, oxygen fluctuations were expected to increase nutrient releases from sediments in comparison with aerobic conditions if oxygen uptake was high enough to produce aerobic-anaerobic fluctuations in the water column.

Materials and methods

Study site presentation, sediment collection and preparation

The study site is Puyvalador reservoir, which is regularly impacted by blooms of cyanobacteria (e.g., Synechococcus sp. and Anabaena spiroides) during the summer. This reservoir is located in the French Pyrenees Mountains (42°38'N, 2°07'E) at altitude of 1421 m. At its maximum storage level, its surface is 90 ha for a volume of 10.1 hm 3 . The maximum depth is 21 m and the mean water residence time is 40 days. This reservoir is part of a hydroelectric complex operated by Electricité de France (EDF) and has two mains tributaries: Aude river and Galbe river. Puyvalador reservoir is eutrophic according to EU Water Framework Directive criteria (Agence de l'eau Rhône Méditerranée Corse, 2015). The central area of the reservoir of Puyvalador, where the depth is maximal, is the main sedimentation zone containing a high percentage (80 %) of fine particles (<63µm) compared to areas influenced by the tributaries (Gautreau et al. in prep). Sediments from the central area are characterized by concentrations of 45 g. kg -1 DW (dry weight) of total organic carbon, 6 g kg -1 DW of total N and 0.5 g kg -1 DW of total P, 33 g.kg -1 DW of Fe, 342 mg.kg -1 DW of Mn, 4 g.kg -1 DW of Ca, and 34 g.kg -1 DW of Al. Measurements of benthic fauna densities during summer period indicated that 4 th stage larvae of Chironomus plumosus and tubificid worms were the most abundant taxa with average densities of 2,095 individuals.m -² and 18,190 individuals.m -², respectively.

For the present experiment, sediments from the central area were collected with an Eckman grab in September 2017. After sampling, they were frozen at -18°C for several weeks to kill the ambient fauna [START_REF] Biles | Bioturbation, ecosystem functioning and community structure[END_REF]. In parallel, a subsample of unfrozen sediment was kept at 15°C (corresponding to the reservoir temperature at the water-sediment interface during sediment collection) to preserve the microbial activities of the collected sediments and to be used as an inoculum for the experiment. Fifteen days before mesocosm preparation, this subsample was sieved at 80 µm to remove the ambient fauna and added to unfrozen sediments. This incubation time of 15 days was long enough as [START_REF] Mermillod-Blondin | A comparison of two ultrasonic methods for detaching biofilms from natural substrata[END_REF] showed a bacterial recolonization of previously sterilized sediments in less than 1 week.

Mesocosms and experimental set up

Experiments were performed in 15 mesocosms made of Plexiglas cylinders with an internal diameter of 10 cm and a height of 20 cm (Figure 1). Mesocosms were placed in a 15°C temperature-controlled room for mimicking summer temperature at the water-sediment interface of the reservoir and filled with a layer of incubated sediments (10 cm) overlaid with synthetic water (10 cm) (composition of synthetic water: 96 mg.L -1 NaHCO3; 39.4 mg.L -1 CaSO4•2H2O; 60 mg.L -1 MgSO4•7H2O; 4 mg.L -1 KCl; pH = 7.5; [START_REF] Us Epa | Methods for measuring the acute toxicity of effluents and receiving waters to freshwater and marine organisms[END_REF]). Oxic conditions were maintained in the mesocosms by bubbling water column with atmospheric air (Figure 1). Mesocosms were equipped with peristaltic pumps for a continuous recirculation of the water column to avoid DO stratification but water flow was adjusted to limit sediment resuspension (6 mL.min -1 ). Oxygen probes were placed on the recirculation system of peristatic pumps to measure DO concentrations.

Mesocosms were run over a period of 14 days to homogenize the experimental conditions among systems. After this period, we measured oxygen uptake in all mesocosms to make sure that similar aerobic microbial activities were occurring in the sediment of all mesocosms. Afterwards, the water column of all mesocosms was renewed with synthetic water to eliminate the initial flush of nutrients during the filling of mesocosms with sediments and water.

Fig. 1 Schema of mesocosm installation with peristaltic pump, oxygen sensor and air source. Sediment and water column compartments were 10 cm in height, corresponding to a volume of 0.785 L for each compartment.

Experimental design

A total of 15 mesocosms was set up to quantify the influence of fauna and DO on nutrient fluxes. For testing the influence of fauna, the water column of the mesocosms was saturated in DO. Three conditions were tested with three mesocosms per condition: 1) control mesocosms without fauna, 2) mesocosms with addition of 150 tubificid worms (more than 90% belonging to the species Tubifex tubifex, Michel Lafond, pers. com) and 3) mesocosms with addition of 15 4 th stage larvae of Chironomus plumosus. Tested animals were from professional breeders.

Densities (19,230 individuals per m² for tubificid worms and 1,923 individuals per m² for Chironomus plumosus)

fitted with natural densities measured in the field in summer (see Study site presentation). Three conditions with three replicated mesocosms per condition were used to test the influence of DO concentrations on nutrient fluxes: 1) mesocosms with a water column saturated in DO (same systems than those used as control mesocosms without fauna), 2) mesocosms without fauna and experiencing DO decrease every 12 hours (aeration turn on and off every 12 hours), 3) mesocosms without fauna experiencing anaerobic conditions in the water column.

Measurements

Nutrient dynamics and aerobic microbial respiration were measured during the 15 days of experiment to study the effects of oxygen conditions and bioturbation on biogeochemical processes. Nutrient dynamics were measured by collecting water samples from the water column at 7 times: on Day 0 just before the treatment application and on days 3, 6, 8, 10, 13 and 15 after the treatment application. At each sampling date, a volume of 30 mL of overlying water was sampled from the top of the water column and filtered through a GF/F filters (0.7 µm, Whatman) and conserved at 4°C before chemical analyses. N-NH4 + , N-NO3 -, P-PO4 3-and SiO2 concentrations of all samples were determined by colorimetric methods with a sequential analyzer (SmartChem200, AMS Alliance). During the experiment, the sampled volume of water for analyses (30 mL) was replaced with 30 mL of artificial water for keeping a constant volume of overlying water. Moreover, to prevent oxygenation of anoxic mesocosms during sampling, collected water was replaced by synthetic water at the top of the mesocosms without air entrance. DO fluxes were measured to evaluate the influence of bioturbating fauna on aerobic microbial respiration. These measurements were performed a day before fauna introduction (day -1) and on days 2, 7 and 14 after the addition of fauna. At each sampling date, mesocosms were sealed for avoiding contact with the atmosphere and the incubation time was limited to 4 hours to avoid the decrease of DO concentrations below 5 mg.L -1 . DO was measured with a Pocket Oxygen Meter FireStingGO2 (Pyroscience) every hour during the incubations. DO fluxes in the mesocosms were calculated from the linear decrease of DO over time and expressed as µmol of O2 consumed per hour and m 2 of sediment surface.

Data analysis

For chemical variables and DO fluxes measured at different times (after the treatment application), the effect of fauna was tested using one-way repeated measures analyses of variance (RM-ANOVAs) on days 3, 6, 8, 10, 13 and 15 for nutrients concentrations and 2, 7 and 14 for DO fluxes with fauna treatment (i.e. control without fauna, chironomids, and tubificids treatments) as the main effect. The effect of DO was also tested by one-way RM-ANOVAs on days 3, 6, 10, 13 and 15 for nutrients with oxygenation treatment (i.e. control with saturated DO, DO fluctuation, and anoxia treatments) as the main effect. When one-way RM-ANOVAs were significant (i.e., p<0.05), Tukey's post hoc tests were performed for the last day of the experiment to evaluate which treatments differed after 15 days of mesocosm stabilization. For all variables and before statistical analyses, the normality and the homoscedasticity of the residues were verified using the Shapiro-Wilk's test and the Bartlett's test, respectively. All statistical analyses were done with the R v.3.3.2 software.

Results

Influence of bioturbation on nutrient dynamics, nutrient release rates and DO fluxes Visual observations during the experiment

During the experiment, sediments of control mesocoms presented a dark color characteristics of anoxic conditions (Fig. 2A). Chironomid larvae produced U-shaped tubes in the top 5 cm of the sediment (Fig. 2B). The ventilation of these structures by the chironomids produced a brown colored oxic zone on the tube walls surrounded by dark colored anoxic sediments in the non-bioturbated zones (

Nutrient concentration dynamics and release rates

N-NH4 + concentrations measured in the water column of all treatments significantly increased during the first 6 to 8 days of the experiment and decreased thereafter (Figure 3A, one-way RM-ANOVA, time effect, ***p < 0.0001). This initial N-NH4 + increase was significantly higher in the chironomid treatment compared to the control and tubificid treatments (one-way RM-ANOVA, fauna treatment, *p < 0.01). Nevertheless, after 15 days of mesocosm stabilization with fauna, N-NH4 + concentrations were comparable in control, tubificid worm and chironomid treatments (Tukey's test, p-value > 0.05).

N-NO3 -concentrations measured in the water column significantly increased during the course of the experiment from 0.1 mg.L -1 at day 0 to 1.1, 2.2 and 3.4 mg.L -1 at the end of the experiment for the control, the tubificid and the chironomid treatments, respectively (Figure 3B, one-way RM-ANOVA, time effect, ***p <0.0001). These N-NO3 -increases differed over time depending on the treatment (one-way RM-ANOVA, "fauna treatment * time" effect, ***p <0.0001). Significant N-NO3 -increases occurred on the first days for the controls and in presence of tubificids whereas N-NO3 -concentrations increased between day 6 and day 13 in presence of chironomids (Figure 3B). Such N-NO3 -increase in the chironomid treatment was concomitant with the N-NH4 + decrease of concentration in the water column (Figure 3A). At the end of the experiment, N-NO3 -concentrations were significantly higher in the chironomid treatment compared to the control and tubificid treatments (Tukey's tests, p-value < 0.05). Considering both inorganic N-NH4 + and N-NO3 -, concentrations, inorganic N concentrations measured on day 15 in chironomid and tubificid treatments were 3.7-fold and 2-fold higher than in control treatment, respectively (Tukey's tests, p-value < 0.05). P-PO4 3-concentrations measured in the water column in presence of chironomids significantly increased from 1 µg.L -1 at day 0 to 60 µg.L -1 at day 6 and then decreased to reach 21 µg.L -1 at day 15 (Figure 3C). In comparison, P-PO4 3-concentration in the water column of control and tubificid treatments remained below 10 µg.L -1 during the whole course of the experiment. Then, after 15 days of experiment, P-PO4 3-concentrations in chironomid treatment was 17-fold higher than those measured in control treatment. Despite low P-PO4 3-concentrations, the occurrence of tubificid worms also increased by 3-fold P-PO4 3-concentrations in water column in comparison with control treatment (Tukey test, p-value < 0.05).

SiO2 concentrations measured in the water column significantly increased in all treatments during the course of the experiment (Figure 3D, one-way RM-ANOVA, time effect, ***p < 0.0001) and reached 20 mg.L -1 for the chironomid treatment and 13 mg.L -1 for the control and tubificid treatments at day 15 (one-way RM-ANOVA, treatment effect, *p < 0.01). Such SiO2 increase was significantly higher in the chironomid treatment than in other treatments (Tukey's tests, *p < 0.01). At day 15, concentration of SiO2 was 1.4-fold higher in chironomid treatment than in control treatment.

Oxygen uptake DO uptake rates were comparable in all mesocosms before the fauna addition (around 1500 µmol.h -1 .m -2 ) and significantly varied during the course of the experiment with the highest rates measured on day 7 (Figure 4, oneway RM-ANOVA, time effect, ***p < 0.0001). After the fauna addition, the presence of chironomid larvae significantly increased oxygen uptake rates in comparison with the other treatments (one-way RM-ANOVA, treatment effect, ***p < 0.0001; Tukey's tests, p *< 0.01). For example, mean DO uptake rate measured on day 14 was 2-fold higher in the chironomid treatment (i.e. 2665 µmol.hour -1 .m -2 ) than in the control treatment (i.e. 10 1196 µmol.hour -1 .m -2 ). Mean DO uptake rates measured in tubificid treatment were slightly higher than in control treatment but this effect was not statistically significant at all dates (see post-hoc Tukey's tests in Figure 4).

Influence of oxygen conditions in the water column on nutrient dynamics and release rates

Our experimental approach allowed generating contrasted oxic/anoxic conditions in the water column for the three treatments. DO concentrations were maintained above 9 mg/L in the water column for the DO-saturated control treatment whereas they were below 0.05 mg/L in the anoxic treatment during the whole experiment. In the DO fluctuation treatment, DO concentrations varied daily from 9 mg/L during the water column aeration to 5 mg/L about 12 h after the shutdown of the water column aeration (Supplementary Material 1). N-NH4 + concentration measured in the water column significantly increased from 0.5 mg.L -1 to 6.4 mg.L -1 during the course of the experiment in the anoxia treatment whereas only small variations in N-NH4 + concentrations were measured in the DO-saturated control and the DO fluctuation treatments (Figure 5A, one-way RM-ANOVA, 'treatment * time" effect, ***p < 0.0001). Indeed, at the end of the experiment, concentration of N-NH4 + was 102-fold time higher in anoxia treatment than in DO-saturated control treatment (Tukey's tests,*p < 0.01).

In contrast, N-NO3 -concentrations in the water column significantly increased in the DO-saturated control and DO-fluctuation treatments from 0 mg.L -1 on day 0 to about 1 mg.L -1 from day 3 to day 15 whereas N-NO3 - concentrations in the anoxia treatment remained low (around 0.04 mg.L -1 ) throughout the experiment (Figure 5B, one-way RM-ANOVA, "treatment * time" effect, ***p < 0.0001). Then, N-NO3 -concentration was 27-fold time higher in control treatment than in anoxia treatment at the end of the experiment (Tukey's tests, *p < 0.01).

Similarly to N-NH4 + dynamics, P-PO4 3-concentrations in the water column significantly increased from 0 to 1034 µg.L -1 during the course of the experiment in the anoxia treatment whereas P-PO4 3-concentrations remained low (around 3 µg.L -1 ) in the two other treatments (Figure 5C, one-way RM-ANOVA, "treatment * time" effect, ***p < 0.0001). At the end of the experiment, concentration of P-PO4 3-in the water column was 56-fold time higher in anoxia treatment than in control treatment (Tukey's tests, *p < 0.01). SiO2 concentration measured in the water column significantly increased in all treatments during the course of the experiment (Figure 5D, one-way RM-ANOVA, time effect, ***p < 0.0001) with higher increases detected in the anoxia treatment (e.g., 19.2 mg.L -1 at day 15) compared to the other treatments (e.g., about 13-14 mg.L -1 at day15) (one-way RM-ANOVA, treatment effect, *p < 0.01). * Concentrations of P-PO4 3-in the anoxia treatment at day 15 are lacking due to an analytical problem.

Discussion

Influence of fauna on the dynamics of nutrients in water column

Nutrient concentrations in water column were highly variable during the course of the experiment in fauna treatments. More precisely, the occurrence of chironomids in mesocosms produced peaks of N-NH4 + and P-PO4 3- in the water column during the first 8 days of experiment. These releases of reduced compounds (N-NH4 + and P-PO4 3-) from the pore water of sediments were more probably due to the initial installation (building and ventilation of U-shaped tubes) of chironomids in poorly-oxygenated sediments (as illustrated in Figure 6). This effect of chironomid larvae on the flush of reduced compounds from sediments to the water column has been commonly observed in experimental studies with organic-rich lake sediments [START_REF] Pelegri | Nitrogen cycling in lake sediments bioturbated by Chironomus plumosus larvae, under different degrees of oxygenation[END_REF][START_REF] Hansen | The impact of Chironomus plumosus larvae on organic matter decay and nutrient (N, P) exchange in a shallow eutrophic lake sediment following a phytoplankton sedimentation[END_REF][START_REF] Stief | Probing the microenvironment of freshwater sediment macrofauna: Implications of deposit-feeding and bioirrigation for nitrogen cycling[END_REF]. Nevertheless, this effect was temporary in the present experiment and, after an initial flush, N-NH4 + concentrations decreased concomitantly with an increase of N-NO3 - concentrations in the water column. This evolution of inorganic N compounds in the water column corresponds to a classic sequence of nitrification [START_REF] Bowen | Connecting the dots: linking nitrogen cycle gene expression to nitrogen fluxes in marine sediment mesocosms[END_REF][START_REF] Pigneret | Urban pollution of sediments: Impact on the physiology and burrowing activity of tubificid worms and consequences on biogeochemical processes[END_REF], which was more likely performed by aerobic microorganisms present in the oxygen-saturated water column and the top oxidized sediment layer.

Since this process was mainly detected in presence of chironomids, we also expect that the burrow walls acted as a very active zone for nitrification. Several authors demonstrated that burrow construction and ventilation by Chironomus plumosus could increase the aerobic surface of the water-sediment interface, and then the favorable zones for nitrification and coupled nitrification-denitrification [START_REF] Pelegri | Nitrogen cycling in lake sediments bioturbated by Chironomus plumosus larvae, under different degrees of oxygenation[END_REF][START_REF] Svensson | Effects of bioturbation by tube-dwelling chironomid larvae on oxygen uptake and denitrification in eutrophic lake sediments[END_REF][START_REF] Lewandowski | Effect of macrozoobenthos on two-dimensional small-scale heterogeneity of pore water phosphorus concentrations in lake sediments: A laboratory study[END_REF][START_REF] Lewandowski | The relationship between Chironomus plumosus burrows and the spatial distribution of pore-water phosphate, iron and ammonium in lake sediments[END_REF][START_REF] Moraes | Nitrification and denitrification in estuarine sediments with tube-dwelling benthic animals[END_REF]. This effect was illustrated on Figure 2B with oxidized sediment nearby U-shaped tubes. Therefore, after 15 days of stabilization, the influence of chironomid larvae bioturbation on N fluxes was predominantly associated with a release of N-NO3 -from oxidized sediments (Figure 6B) than a release of N-NH4 + from poorly-oxygenized sediments (as observed during chironomid installation, Figure 6A). Concerning P dynamics, the presence of chironomids induced an initial flush of P-PO4 3-from the sediment to the water column (i.e, up to 60 mg.L -1 ) during the first week of the experiment followed by a small decrease of P-PO4 3-concentrations in the water column the second week of the experimentation, which remained around 30-40 µg.L -1 until the end of the experiment. Then, the stimulating effect of chironomid ventilation on P-PO4 3-concentrations compared to the control treatment without fauna was detected until the end of the experiment. According to [START_REF] Lewandowski | Effect of macrozoobenthos on two-dimensional small-scale heterogeneity of pore water phosphorus concentrations in lake sediments: A laboratory study[END_REF] and [START_REF] Hupfer | Chironomid larvae enhance phosphorus burial in lake sediments: Insights from long-term and short-term experiments[END_REF], the supply of DO in chironomid burrows could have produced an adsorption of the P-PO4 3-initially flushed in the water column into Fe(OH)3-PO4 complexes in Fe-rich sediments (ie. 33 g.kg -1 DW of total Fe, see materials and methods). Such mechanism probably occurred during our experiment inducing the small decrease of P-PO4 3-in the water column after day 6 but did not lead to a complete adsorption of P-PO4 3-flushed from the sediment. We suppose that P-PO4 3-adsorption occurring in aerobic conditions and enhanced by chironomid ventilation was compensated by a constant flush of P-PO4 3-from the anoxic sediments due to chironomid bioturbation (Figure 6). As the dynamics of nutrients during the course of the experiment in chironomid treatment were largely associated with the installation of benthic chironomid larvae in sediments, we compared the effects of three fauna treatments on nutrients in water column by considering the last day of the experiment after 15 days of stabilization of biogeochemical processes in the mesocosms.

Influence of bioturbation on nutrient concentrations and oxygen uptake

According to our prediction, the ventilation of U-shaped burrows by Chironomus plumosus highly increased inorganic N (x 3.7) and P-PO4 3-(x 17) concentration in water column after 15 days of experiment by creating water exchanges at the water sediment interface. Indeed, the presence of oxic brown sediments observed on the burrow walls of chironomids demonstrated a transport of oxygen-saturated water from the water column in the biogenic structures built by animals in anoxic sediments. This ventilation of U-shaped burrows produced a significant stimulation of oxygen uptake (2-fold increase on day 7 compared to control treatment without fauna) by increasing the supply of DO in sediments. This stimulation of oxygen uptake by chironomids probably contributed to nutrient dynamics by stimulating microbial activities and organic matter mineralization in sediments [START_REF] Aller | Transport and reaction in the bioirrigated zone[END_REF].

Nevertheless, the major mechanism explaining the influence of chironomid larvae on nutrient fluxes at the water-sediment interface was burrow ventilation which increased the release rates of nutrients (inorganic nitrogen N-NH4 + + N-NO3 --and P-PO4 3-) accumulated in pore water [START_REF] Lewandowski | Effect of macrozoobenthos on two-dimensional small-scale heterogeneity of pore water phosphorus concentrations in lake sediments: A laboratory study[END_REF][START_REF] Lewandowski | The relationship between Chironomus plumosus burrows and the spatial distribution of pore-water phosphate, iron and ammonium in lake sediments[END_REF]. The increase (+ 50%) of SiO2 concentration in water column with Chironomus plumosus was in accordance with this mechanism of burrow ventilation as silicate fluxes were found to depend on irrigation of biogenic structures by benthic fauna (e.g., [START_REF] Mermillod-Blondin | Influence of bioturbation by three benthic infaunal species on microbial communities and biogeochemical processes in marine sediment[END_REF].

In contrast with chironomid larvae, tubificid worms had lower influences on N (x2 in comparison with the control treatment) and P (x3 in comparison with the control treatment) concentrations in the water column in our experimental conditions. Such contrasted effects of Chironomus plumosus and tubificid worms on nutrient concentrations were related to their different bioturbation activities in the sediment. Tubificid worms produce deep biogenic structures in the sediment column but they do not ventilate their galleries to the same extent as chironomids [START_REF] Leuchs | Die Schlaengelaktivität von Chironomuslarven (Diptera) aus flachen und tiefen Gewässern und die resultierenden Wasserzirkulationen in Abhängigkeit von Temperatur und Sauerstoffangebot[END_REF][START_REF] Wood | Role of oligochaetes in the circulation of water and solutes across the mud-water interface[END_REF][START_REF] Svensson | Nitrification and denitrification in a eutrophic lake sediment bioturbated by oligochaetes[END_REF]. In presence of tubificid worms, SiO2 concentrations were not significantly different compared to the controls without fauna, indicating that tubificids did not strongly enhance water exchanges, oxygen uptake and release of reduced compounds (N-NH4 + and P-PO4 3- ) at the water-sediment interface (Figure 7). Our results contrasted with several studies showing a significant stimulating effect of tubificid worms on oxygen uptake at the water sediment interface [START_REF] Fukuhara | Enhancement of inorganic nitrogen and phosphate release from lake sediment by tubificid worms and chironomid larvae[END_REF][START_REF] Mermillod-Blondin | Ecosystem engineering: the impact of bioturbation on biogeochemical processes in marine and freshwater benthic habitats[END_REF][START_REF] Saaltink | Respiration and aeration by bioturbating Tubificidae alter biogeochemical processes in aquatic sediment[END_REF]. For example, [START_REF] Mermillod-Blondin | Laboratory study highlights the key influences of stormwater sediment thickness and bioturbation by tubificid worms on dynamics of nutrients and pollutants in stormwater retention systems[END_REF] showed that tubificid worms at a density (21,000 individuals.m -2 ) close to those used in our experiment (19,230 individuals.m -2 ) stimulated by around 2-fold oxygen uptake from sediments rich in organic matter. Such contrasting results were more likely due to differences in sediment characteristics, which could largely modulate the functional significance of bioturbation at the water-sediment interface (e.g. [START_REF] Nogaro | Alum application, invertebrate bioturbation, and sediment characteristics interact to affect phosphorus exchange in eutrophic ecosystems[END_REF]. In the present study, total organic carbon concentrations in sediments were indeed 4-fold lower than concentrations in sediments tested by [START_REF] Mermillod-Blondin | Laboratory study highlights the key influences of stormwater sediment thickness and bioturbation by tubificid worms on dynamics of nutrients and pollutants in stormwater retention systems[END_REF] (i.e. about 4.5% vs. 20% DW, respectively). Sediment organic content could have a significant influence on nutrient dynamics because the microbial mineralization of organic matter in the sediment determines (1) the production of inorganic nutrients (N-NH4 + and P-PO4 3-) and (2) the redox conditions in the sediment column [START_REF] Forsberg | Importance of sediments in understanding nutrient cyclings in lakes[END_REF][START_REF] Kristensen | Organic matter diagenesis at the oxic/anoxic interface in coastal marine sediments, with emphasis on the role of burrowing animals[END_REF]. Lower sediment organic content in our study might have induced lower microbial activity and associated redox conditions (eg, more oxygen, less anaerobic sediment processes) compared to [START_REF] Mermillod-Blondin | Laboratory study highlights the key influences of stormwater sediment thickness and bioturbation by tubificid worms on dynamics of nutrients and pollutants in stormwater retention systems[END_REF]. This could explain why N-NO3 - released from sediments was higher in the present study than N-NH4 + released in presence of tubificids compared to control treatment (Figure 3) whereas N-NH4 + concentrations dominated inorganic N fluxes in [START_REF] Mermillod-Blondin | Laboratory study highlights the key influences of stormwater sediment thickness and bioturbation by tubificid worms on dynamics of nutrients and pollutants in stormwater retention systems[END_REF]. Therefore, we hypothesize that the role of bioturbation in nutrient dynamics is tightly linked to the sediment organic content (and associated redox conditions) but further experiments using a wide variety of sediments are needed to clarify this link.

Influence of the oxygenation treatment on nutrient fluxes

According to our prediction, the anoxia treatment increased by 102-fold N-NH4 + concentrations and by 56-fold P-PO4 3-concentrations in comparison with the well-oxygenated control treatment. In contrast, the treatment with DO-fluctuations did not affect nutrient fluxes in comparison with the control treatment because sediment DO uptake was not high enough to produce anoxic conditions in the water column (DO concentrations never dropped below 5 mg.L -1 during the experiment). In consequence, DO fluctuations had no significant effect on P released from sediment as desorption of P from Fe, Mn, Al or Ca only occurs under anoxic conditions [START_REF] Dahm | Role of anaerobic zones and processes in stream ecosystem productivity[END_REF][START_REF] Boulton | The functional significance of the hyporheic zone in streams and rivers[END_REF]. Moreover, the Fe:P ratio (by weight) in Puyvalador sediments is around 66 (so above 15) and can control internal P loading in presence of oxygen [START_REF] Jensen | Iron/phosphorus ratio in surface sediment as an indicator of phosphate release from aerobic sediments in shallow lakes[END_REF]. Moreover, DO fluctuations did not stimulate organic matter mineralization in sediments, a mechanism reported by [START_REF] Aller | Transport and reaction in the bioirrigated zone[END_REF] in marine sediments.

Several studies highlighted that anaerobic condition in the water column was a major factor for the release rates of N-NH4 + and P-PO4 3-from sediments [START_REF] Boström | Exchange of phosphorus across the sedimentwater interface[END_REF][START_REF] Hedin | Thermodynamic constraints on nitrogen transformations and other biogeochemical processes at soil-stream interfaces[END_REF]Wetzel, 2001bWetzel, , 2001c;;[START_REF] Beutel | Inhibition of ammonia release from anoxic profundal sediments in lakes using hypolimnetic oxygenation[END_REF][START_REF] Ekeroth | Nutrient fluxes from reduced Baltic Sea sediment: effects of oxygenation and macrobenthos[END_REF][START_REF] Rapin | Phosphorus mobility in dam reservoir affected by redox oscillations: An experimental study[END_REF]. Indeed, N-NH4 + produced by organic matter mineralization was nitrified into N-NO3 -when DO was available for microorganism (in the control and fluctuation treatments) whereas N-NH4 + accumulated in anoxic conditions (in the anoxic treatment). Therefore, fluxes of inorganic N from the sediment to the water column were dominated by N-NO3 -fluxes in the control and fluctuation treatments and by N-NH4 + in the anoxic treatment (Figure 7). P fluxes were linked with adsorption mechanisms modulated by DO concentrations: P-PO4 3-was adsorbed with Fe(OH)3 and Mn(OH)2 in aerobic conditions and released when oxygen was limited (Figure 7). Surprisingly, SiO2 fluxes from sediment to water column were also stimulated by anoxic conditions whereas the SiO2 dynamics was not dependent on redox conditions like P or Fe which precipitated in oxic conditions [START_REF] Tessenow | Lösungs-, diffusions-und sorptionsprozesse in der oberschicht von Seesedimenten[END_REF][START_REF] Granéli | The influence of Chironomus plumosus larvae on the exchange of dissolved substances between sediment and water[END_REF]. We can suppose that this increase of SiO2 fluxes under anoxic conditions was associated with the production and ebullition of CH4 bubbles produced in sediments during the experiment. Indeed, production of gas has been observed in the anoxic treatment during the course of the experiment and it has been shown that the ebullition of CH4 could be as a significant driver of water fluxes and associated dissolved solutes from pore water to water column [START_REF] Klein | Sediment porewater exchange and solute release during ebullition[END_REF][START_REF] Cheng | Diffusion of nitrogen and phosphorus across the sediment-water interface and in seawater at aquaculture areas of Daya Bay, China[END_REF]. Consequently, such process might have contributed to the releases of SiO2 but also reduced nutrients (N-NH4 + and P-PO4 3-) from pore water to water column.

Conclusions

The present study showed that natural densities of Chironomus plumosus larvae could increase N (N-NH4 + + N-NO3 -) and P-PO4 3-release rates by 3.7-and 17-fold, respectively. Anoxic conditions in the water column had higher influences on N (x102 in comparison with the control) and P (x56 in comparison with the control) concentrations in the water column than the most efficient bioturbator -chironomid larvae -(x3.7 for N and x17 for P). These results confirmed the main conclusion of [START_REF] Ekeroth | Nutrient fluxes from reduced Baltic Sea sediment: effects of oxygenation and macrobenthos[END_REF] in marine sediments which stated that oxygen condition was a stronger regulator of P fluxes than benthic fauna. The higher influence of anoxia on N release from sediments in comparison with bioturbation was more surprising and can be due to a lack of denitrification in anoxic conditions. Indeed, the stimulating influence of ventilated U-shaped tubes on nitrificationdenitrification coupling reported in several studies [START_REF] Svensson | Effects of bioturbation by tube-dwelling chironomid larvae on oxygen uptake and denitrification in eutrophic lake sediments[END_REF][START_REF] Moraes | Nitrification and denitrification in estuarine sediments with tube-dwelling benthic animals[END_REF] might have led to N loss from chironomid mesocosms (NH4  NO3  N2) and then a limitation of inorganic dissolved N accumulated in the water column during the course of the experiment. In contrast, the lack of nitrification in anoxic conditions might have suppressed N-NO3 -denitrification and produced a high accumulation of N-NH4 + in the water column during the course of the experiment. Nevertheless, this explanation needs to be verified by measuring the denitrification potential in each experimental treatment.

Overall, we can conclude that the presence of efficient bioturbating fauna and anoxia at the sediment-water interface should not be neglected in the evaluation of nutrient dynamics in lake sediments during the warmer season of the year. This demonstrates the need to integrate environmental biogeochemistry with the dynamics of benthic invertebrates to provide reliable and ecologically relevant assessments of nutrient cycling in reservoirs.

Nevertheless, nutrient dynamics in sediments can be largely dependent on several factors linked to sediment characteristics such as organic matter concentration and Fe content (e.g., [START_REF] Forsberg | Importance of sediments in understanding nutrient cyclings in lakes[END_REF][START_REF] Jensen | Iron/phosphorus ratio in surface sediment as an indicator of phosphate release from aerobic sediments in shallow lakes[END_REF].

Extending our experimental approach to several reservoir sediments is greatly needed to evaluate how the role of

  Fig 2B). Tubificid worms dug galleries in the whole sediment column (10 cm deep, Fig 2C) and deposited fecal pellets at the sediment surface. Nevertheless, we did not observe any brown oxidized zone on the gallery walls produced by the tubificid worms.

Fig. 2

 2 Fig. 2 Pictures from the mesocosms: (A) control treatment without fauna, (B) U-shaped tubes produced by chironomid larvae on the inner wall of a mesocosm and (C) galleries produced by tubificid worms on the inner wall of a mesocosm.

Fig. 3

 3 Fig. 3 Concentrations of N-NH4 + (A), N-NO3 -(B), P-PO4 3-(C), and SiO2 (D) measured in the water column of mesocosms during the course of the experiment for the three fauna treatments (mean ± standard error, n=3).

Fig. 4

 4 Fig. 4 Oxygen uptake rates (fluxes from water column to sediment) measured for the three fauna treatments (mean ± standard error, n=3 mesocosms). For each panel, different letters (A, B and C) show significant differences among treatments (see axis label, Tukey post hoc test, p<0.05).

Fig. 5

 5 Fig. 5 Dynamics of N-NH4 + (A), N-NO3 -(B), P-PO4 3-(C), and SiO2 (D) measured in the water column during the course of the experiment for the three DO treatments (mean ± standard error, n=3).

Fig. 6

 6 Fig. 6 Summary of the main influence of fauna bioturbation processes on water fluxes and associated nutrient fluxes: (A) burrow construction and ventilation by Chironomus plumosus during the first week and (B) during the second week. The arrows from the water to the sediment compartment represent the fluxes of water and DO (larger arrow means larger fluxes). Arrows from the sediment to the water compartment show the release of nutrients.

Fig. 7

 7 Fig. 7 Summary of the main influence of DO conditions in the water column on nutrient fluxes (A) oxic conditions and (B) anoxic conditions.

Fig. 1

 1 Fig.1Schema of mesocosm installation with peristaltic pump, oxygen sensor and air source. Sediment and water column compartments were 10 cm in height, corresponding to a volume of 0.785 L for each compartment.

Fig. 2

 2 Fig. 2 Pictures from the mesocosms: (A) control treatment without fauna, (B) U-shaped tubes produced by chironomid larvae on the inner wall of a mesocosm and (C) galleries produced by tubificid worms on the inner wall of a mesocosm.

Fig. 3

 3 Fig. 3 Concentrations of N-NH4 + (A), N-NO3 -(B), P-PO4 3-(C), and SiO2 (D) measured in the water column of mesocosms during the course of the experiment for the three fauna treatments (mean ± standard error, n=3).

Fig. 4

 4 Fig.4Oxygen uptake rates (fluxes from water column to sediment) measured for the three fauna treatments (mean ± standard error, n=3 mesocosms). For each panel, different letters (a, b and c) show significant differences among treatments (see axis label, Tukey post hoc test, p<0.05).

Fig. 5

 5 Fig. 5 Dynamics of N-NH4 + (A), N-NO3 -(B), P-PO4 3-(C), and SiO2 (D) measured in the water column during the course of the experiment for the three DO treatments (mean ± standard error, n=3).

Fig. 6

 6 Fig. 6 Summary of the main influence of fauna bioturbation processes on water fluxes and associated nutrient fluxes: (A) burrow construction and ventilation by Chironomus plumosus during the first week and (B) during the second week. The arrows from the water to the sediment compartment represent the fluxes of water and DO (larger arrow means larger fluxes). Arrows from the sediment to the water compartment show the release of nutrients.

Fig. 7

 7 Fig. 7 Summary of the main influence of DO conditions in the water column on nutrient fluxes (A) oxic conditions and (B) anoxic conditions.

Fig. 1

 1 Fig. 1 Schema of mesocosm installation with peristaltic pump, oxygen sensor and air source. Sediment and water column compartments were 10 cm in height, corresponding to a volume of 0.785 L for each compartment

Fig. 4

 4 Fig. 4 Oxygen uptake rates (fluxes from water column to sediment) measured for the three fauna treatments (mean ± standard error, n=3 mesocosms). For each panel, different letters (a, b and c) show significant differences among treatments (see axis label, Tukey post hoc test, p<0.05).

Fig. 5 Fig. 6

 56 Fig. 5 Dynamics of N-NH4 + (A), N-NO3 -(B), P-PO4 3-(C), and SiO2 (D) measured in the water column during the course of the experiment for the three DO treatments (mean ± standard error, n=3).

Fig. 7

 7 Fig. 7 Summary of the main influence of DO conditions in the water column on nutrient fluxes (A) oxic conditions and (B) anoxic conditions.

Acknowledgments

The authors thank the ANRT (Association Nationale Recherche Technologie) and EDF (Electricité de France) for Edwige Gautreau's Ph.D. grant (convention bourse CIFRE-EDF: n° 0733/2016). This study was funded by a

Funding

This study was funded by the partnership EDF-CNRS (CNRS n°149389) and Lyon Metropole within the framework of Sedaqua Platform (Contrat Plan Etat Région).

bioturbation and anoxia on nutrient dynamics depends on sedimentary characteristics. This is a necessary step to develop a more generalized framework on the contribution of environmental factors on nutrient dynamics in reservoirs during the warm season that may be subject to algal blooms.

Supplementary material

Online Resource 1 (A) Concentration of dissolved oxygen (mg.L -1 ) in control treatment from day 2 to day 14, (B) concentration of dissolved oxygen (mg.L -1 ) in anoxia treatment from day 2 to day 14 and (C) concentration of dissolved oxygen (mg.L -1 ) monitored in fluctuation treatment from day 12 to day 15.

Author's contributions

All authors contributed to the study conception and design. Material preparation, data collection and analysis were performed by Edwige Gautreau, Florian Mermillod-Blondin and Laurence Volatier. The first draft of the manuscript was written by Edwige Gautreau, Florian Mermillod-Blondin and Laurence Volatier. All authors contributed critically to the drafts and gave final approval for publication.

Ethical approval

All applicable international, national, and/or institutional guidelines for the care and use of animals were followed.

Conflict of Interest

The authors declare that they have no conflict of interest.

Online Resource 1 (A) Concentration of dissolved oxygen (mg.L -1 ) in control treatment from day 2 to day 14, (B) concentration of dissolved oxygen (mg.L -1 ) in anoxia treatment from day 2 to day 14 and (C) concentration of dissolved oxygen (mg.L -1 ) monitored in fluctuation treatment from day 12 to day 15.