# Sediment implication in shift in nutrient limitation and impact on purification capacity: Application on a common Atlantic east coast pond

Raphaël Moncelon, Charlotte de Lignières, Philippe Pineau, Claire Emery, Eric Bénéteau, Edouard Metzger, Christine Dupuy

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## ► To cite this version:

Raphaël Moncelon, Charlotte de Lignières, Philippe Pineau, Claire Emery, Eric Bénéteau, et al. Sediment implication in shift in nutrient limitation and impact on purification capacity: Application on a common Atlantic east coast pond. Science of the Total Environment, 2025, 975, pp.179308. 10.1016/j.scitotenv.2025.179308. hal-05061595

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Submitted on 9 May 2025

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# Sediment implication in shift in nutrient limitation and impact on purification capacity: Application on a common Atlantic east coast pond

<figure><img src="image_4.png" /></figure>

Raphaël Moncelon<sup>a, *</sup>, Charlotte de Lignières<sup>a</sup>, Philippe Pineau<sup>a</sup>, Claire Emery<sup>a</sup>, Eric Bénéteau<sup>b</sup>, Edouard Metzger<sup>b</sup>, Christine Dupuy<sup>a</sup>

## HIGHLIGHTS

* Shift from P to N limitation occurred during the spring/summer transition

* SRP effluxes from the sediment coincided with N-limitation in the water column

* Iron oxide reduction in the sediment caused SRP efflux toward the water column

* SRP efflux significantly contributed to the water enrichment and primary production

* Water residence time altered phytoplankton development and nutrient epuration

## GRAPHICAL ABSTRACT

<figure><img src="image_5.png" /></figure>

## ARTICLE INFO

Editor: Jay Gan

**Keywords:**

Benthos-pelagos coupling

Primary production

Sedimentary input

Nutrient mitigation

## ABSTRACT

Salt marshes are widely recognized as regulation areas of nutrient flows to the ocean. However, poor knowledge of nutrient dynamics in these environments counteracts the understanding of nutrient regulation within the land-sea continuum, previous works mainly focusing on the water compartment. To overcome this gap, an experimental pond was monthly (open system) and weekly (closed system) monitored between February and October 2021. Bio(geo)chemical parameters (in water and sediment) and nutrient fluxes at the sediment-water interface (SWI) were relevantly explored to highlight the coupling between the benthic and pelagic compartments. A significant shift from P limitation (February–June period) to N limitation (July–October period) was observed and attributed to the phosphate mobilization (named soluble reactive phosphorus, SRP) processes in the sediment and its effluxes toward the water column. This SRP mobilization was strongly linked to the respiration of iron oxides for mineralization of organic matter (OM) and anammox processes. Surprising high carbon production in September (>1200 mgC m⁻³ d⁻¹) in a low nutrient concentration range indicated significant re-generated N production and potentially new P production. Monitoring in enclosed system indicated a 3-week water retention time for a maximum nutrient removal and minimal impact for water export to the coastal waters. In non-bloom period, sediment SRP efflux from the sediment could contribute up to 40 % of SRP increase in the water column in September and to 60 % of the P-production demand in October. This work highlighted the importance of sediment for water nutrient re-supply and the temporality of water retention time in pond system

<footnote><sup>a</sup> Laboratoire LIENSs, UMR 7266, La Rochelle Université, Bâtiment ILE, 2 Rue Olympede Gouges, La Rochelle, France</footnote>

<footnote><sup>b</sup> Laboratoire de Planétologie et Géosciences UMR 6112, CNRS, Université d'Angers, Nantes Université, Le Mans Université, Angers, France</footnote>

<footnote>* Corresponding author. LIENSs Laboratory, France.</footnote>

<footnote>E-mail address: raphael.moncelon1@univ-lr.fr (R. Moncelon).</footnote>
to optimize purification effect and limit risks of eutrophication in coastal waters. While the eutrophication potential associated with nitrogen seemed to be reduced in such pond, that of phosphorus needs to be more considered.

# 1. Introduction

The availability of nutrients, particularly nitrogen (N), phosphorus (P) and silicon (Si), is a major issue in the dynamics of primary production, particularly that of phytoplankton (Redfield, 1958; Justic et al., 1995). This can have an impact on phytoplankton biomass and species composition, ecosystem services (Cadée and Hegeman, 1986; Riegman et al., 1992; Humborg et al., 1997; Jickell, 1998) and eutrophication risks (Devlin and Brodie, 2023).

It is commonly established that terrestrial freshwaters are generally soluble reactive phosphorus (SRP) limited (Schindler, 1977; Hecky and Kilham, 1988, Fennel and Testa, 2019) and coastal waters limited in dissolved inorganic nitrogen (DIN) (Ryther and Dunstan, 1971; Hecky and Kilham, 1988; Howarth and Marino, 2006; Paerl, 2018). The latter is partially justified by the excess of SRP input from the sediment (Caraco et al., 1989, 1990), lying on the ability of SRP to sorb and desorb metal oxides in early stage of organic matter (OM) mineralization (Carignan and Flett, 1981; Cornwell and Morse, 1987; Chambers and Odum, 1990; Koop-Jakobsen and Giblin, 2010).

As a consequence of the global positioning of N limitation in saline environments, environmental and scientific interests in the N cycle and its removal processes have intensified for coastal ecosystems (National Research Council, 2000) and in particular in tidal marshes. This is part of the drive to reduce eutrophication in coastal areas (Teal and Howes, 2002; Valiela and Cole, 2002; Fisher and Acreman, 2004). However, increasing population densities and activities in coastal areas disproportionately modified N, P and Si loadings to rivers (Crossland et al., 2005). This may change the type of nutrient limitation for phytoplankton in coastal receiving waters (Falkowski, 2000; Yin et al., 2001) and even leading to co-limitations in both fresh and salt systems (Elser et al., 2007; Maberly et al., 2020; Browning and Moore, 2023). Such nutrient loads from land can even place some coastal waters in P limitation, which can be the case of Bay of Biscay (Labry et al., 2002; Loyer et al., 2006) in which this study fits.

While the evidence for shifts in nutrient limitation in the spatial dimension of the land-sea continuum is widely established (Crain, 2007), the evidence for such shifts within the same environment is less common in the literature (Fisher et al., 1999). Some work in transient waters, such as estuaries where fresh and saltwater mix, showed spatial and temporal changes in the relative availability of N and P causing shifts in nutrient limitation (e. g. Doering et al., 1995; Fisher et al., 1999). In salt marshes, it has been shown that nutrient limitation shifted across estuarine salinity gradients, from N-limitation in salt and brackish marsh system to co-limited by N and P in oligohaline system (Crain, 2007). Even though salt marshes are widely recognized as regulation areas of nutrient flows to the ocean, these changes make difficult the prioritization of management efforts toward particular nutrients (Conley, 2000). Moreover, poor knowledge of nutrient dynamics in these environments counteracts the understanding of nutrient regulation within the land-sea continuum, previous works mainly focusing on the water compartment.

The study of nutrient dynamics in ponds is of particular interest because of their position at the land-sea interface and their key role in the storage and removal of nutrients and pollutants (Risser, 1995; Levin et al., 2001). As ecotone systems, material and nutrient flows and processing are concentrated or even intensified (Traut, 2005) depending on the strength of interactions between adjacent environments (Holland, 1988). They may however have their own physical and biogeochemical characteristics (Johnston, 1993; Risser, 1995; James and Zedler, 2000). However, the impact of the latter on the retroactive dynamics of

nutrients and phytoplankton does not seem to be fully identified. Consequently, the exploitation of the purifying capacities of these environments is not always mastered. This could be caused by the lack of consideration of the sedimentary and aquatic compartment as a whole. Considered but rarely highlighted in the experiments, biogeochemical processes arising from the benthos-pelagos coupling seem to be a determining factor in understanding the purifying functions of wetlands, particularly in marsh areas (Moncelon et al., 2021, 2022).

This work tends to identify the factors contributing to the nutrient availability for primary production in a widespread temperate anthropized salt marsh (pond of the French Atlantic East coast). This was led through the study of nutrient dynamics in the sediment and biochemical parameters of the water columns. A coupled seasonal monitoring of the filling water masses (monthly monitoring between February and October 2021) and of the water retained in the pond (weekly monitored in a closed system over 4 weeks in spring and late summer 2021) enabled to evaluate the effect of the water residence time on the biogeochemical dynamics and purifying functions of the pond. The better understanding of processes controlling nutrient purification in salt marsh will allow appropriate management methods to limit the risks of eutrophication in coastal zones. This work notably highlights the importance of benthic compartment of marsh-type systems for the understanding of nutrient flows within the land-sea continuum of highly anthropized regions.

# 2. Material and methods

## 2.1. Study sites

Monitoring was carried out at L'Houmeau salt marshes site, covering a surface area of 7 ha and comprising 16 ponds of varying sizes. Each of these ponds was supplied with salt water via channels connected to the Pertuis Charentais sea. The monitoring focused on a single 180 m² pond, chosen for its low calcic sediment composition (shells) that would have interfered with sediment analysis (Fig. S1, supplementary materials).

## 2.2. Sampling strategy and parameters studied

The water column was monitored monthly from February to October 2021. The water in the pond was renewed once a month (tide coefficient > 70) and at least three days before the start of the manipulations, by means of a valve and pipe system. Prior to each field measurement, the water level in the pond was set at around 90 cm. Two weekly follow-ups were realized in spring (between May and June) and late summer (September), each covering 4 weeks (annotated as Spring Sᵢ and Sept Sᵢ for spring and late summer monitoring respectively, i the number of the week). During these monitorings, no water renewal was realized (a vertical pipe was placed at the basin's filling mouth to prevent water intake even at high coefficients). The sediment compartment was sampled every third month in February, April, May, June, August and October. When planned in the same month, the water (monthly monitoring) and sediment were sampled on the same day.

Discrete samples, whether abiotic or biotic, described below were realized in triplicates. For each replicate, water sampling (for phytoplankton and nutrient analysis) was carried out from a water pooling of 15 L from three different zones of the pond.
## 2.3. Abiotic parameters

### 2.3.1. Dissolved phase

In situ temperature, salinity, dissolved oxygen (DO) concentration/ saturation (HOBO U-26 sensor), turbidity (STBD sensor) and pH in the water column were measured continuously between February and October 2021, and/or monthly according to water sampling mentioned above. Nitrate ($NO_3$), nitrite ($NO_2$), ammonium ($\sum NH_3$) (mentioned as DIN), soluble reactive phosphorus (SRP) and silica ($Si_d$) concentrations in water column and pore water samples (see below for porewater extraction) were determined by a colorimetric autoanalyzer Scalar Seal. Moreover, alkalinity was estimated colorimetrically in porewater according to Sarazin et al. (1999) and dissolved iron ($Fe_d$), manganese ($Mn_d$) and sulphur ($S_d$) were measured by ICP-AES (more details in Moncelon et al. (2022)).

### 2.3.2. Solid phase

At each sediment campaign, sediment was collected through three 10 (Ø)*30(H)cm PVC cores and conditioned with a 0.5 cm resolution slicing method. Sediment slicing was realized under nitrogen to avoid oxidation of solid and dissolved phases (more details in Moncelon et al. (2021)). Porewaters were extracted and filtered using a 0.2 µm RC25 Sartorius filter for analysis described in 2.3.1).

Remaining sediment was freeze-dried for porosity calculation and manually grounded for ascorbate extraction of reactive manganese ($Mn^{3+}$ and $Mn^{4+}$, mentioned as $Mn_{asc}$), iron (amorphous $Fe^{3+}$, mentioned as $Fe_{asc}$) and P according to the method used in Kostka and Luther (1994) and Anschutz et al. (1998) (more details in Moncelon et al. (2021)). An aliquot (~15 g) was used for organic matter (OM) quantification by loss on ignition at 430 °C (Davies, 1974). Data are expressed in percentage of dry weight (dw %).

Total chlorophyll a (Chla) biomass and pheo-pigment (Pheo) was measured on freeze-dried sediment by fluorimetry according to the Yentsch and Menzel (1963) method (more details in Moncelon et al. (2021)). This biomass was used as proxy of phytobenthos. The relative proportion of Chla to the amounts of Chla+Pheo was used to estimate the vitality (mentioned as $Chla_{act}$, %) of phytoplankton cells.

### 2.3.3. Diffusive fluxes modeling

$NO_3$, $NO_2$, SRP, $\sum NH_3$ and Si diffusive fluxes (F) at the SWI were determined using the Profile software (Berg et al., 1998) based on the First Fick's Law calculation as:

$$F = - \phi Ds (dC/dz) \quad (1)$$

with Ø the sediment porosity, Ds the element diffusion coefficient in sediment ($D_s = D/(1 + 3(1-\phi))$), according to Rasmussen and Jorgensen (1992)), C the element pore water concentration and z the sampling depth (more detailed in Moncelon et al. (2021, 2022)). In this work, the negative flux values corresponded to fluxes leaving the sediment, the positive values to fluxes entering the sediment.

## 2.4. Biotic parameters of the water column

### 2.4.1. Phytoplankton compartment

Chla (and Pheo-pigment) biomass was evaluated for three-size classes of phytoplankton (micro-: >20 µm, nano-: 3 to 20 µm and picophytoplankton: <3 µm) by fluorimetry according to the Yentsch and Menzel (1963) method (more details in Moncelon et al. (2021)).

Carbon primary production (Prod) measurements ($^{14}$C) by size classes was carried out according to Nielsen (1951) method (more details in Moncelon et al. (2022)).

### 2.4.2. Mesozooplankton compartment

For each campaign, a 200 µm mesh size net was deployed over a total length of approximately 25 m. Samples were preserved in ethanol 70 %

before being identified and counted rapidly under a binocular loupe. The major freshwater groups were counted, i. e. copepods (plus copepodit and nauplii), cladocerans, ostracods, insect larvae, mite, oligochaete larvae and rotifers.

## 3. Results

### 3.1. Water column

#### 3.1.1. Integrated physical-chemical parameters

Water temperature increased from <10 °C (winter and fall) to above 25 °C during spring and summer period, with day/night fluctuations up to 10 °C (Fig. 1, Table 1).

Salinity globally increased from March to September from 27 to 39 ‰ respectively before decrease in October between 29 and 31 ‰ (Fig. 1, Table 1). Mean DO concentration and saturation fluctuated around 9 mg L⁻¹ and above 100 % respectively (Fig. 1). Episodic anoxic events were observed between April and the end of August. Maxima were observed during the late summer weekly monitoring (18 mg L⁻¹ and 250 % respectively) and around mid-October (22 mg L⁻¹ and 270 % respectively). DO values tended to decrease during the spring monitoring (minima in S4 at 4.26 mg L⁻¹). Turbidity showed great amplitude variations from almost 0 to >300 NTU all along the monitoring (Fig. 1) and pH varied between 7.97 (May) and 8.75 (September, S3) (Table 1).

#### 3.1.2. Nutrient dynamics

Nutrient fluctuations in the water column globally followed each other over along monthly and weekly spring monitoring sampling period (Fig. 2). $NO_3$ varied from 37.0 ± 1.0 µmol L⁻¹ (February) to almost 0 µmol L⁻¹ (June) during monthly monitoring (Fig. 2 A) with lower mean concentration from May to October. It decreased during the spring monitoring, from almost 3 µmol L⁻¹ (Spring S1) to 0.5 µmol L⁻¹ (Spring S2, S3, S4) (Fig. 2 B). $NO_2$ fluctuated monthly between 0.1 (September) and 2.2 ± 1.3 µmol L⁻¹ (July) and $\sum NH_3$ between almost 0 and 7.0 ± 0.5 µmol L⁻¹ in June and October respectively. During the spring monitoring, $NO_2$ and $\sum NH_3$ both decreased between S1 and S3 and re-increased in S4. The late summer weekly monitoring showed mor erratic fluctuation in DIN concentration, with same trend between $NO_3$ and $\sum NH_3$ (maximum concentration in S2 and S4), whereas $NO_2$ peaked in S3 (Fig. 2 C).

SRP varied between 0.1 (March) and 0.9 µmol L⁻¹ (February) between February and June in the monthly monitoring. The concentrations were higher from July to October, between 0.8 (September) and 1.4 µmol L⁻¹ (August and October) (Fig. 2 A). As for DIN, SRP decreased between Spring S1 and S3 before re-increase in S4 (Fig. 2 B). Same trend was observed in the late summer weekly monitoring with stronger depletion in S2 and S3 (Fig. 2 C).

Finally, Si fluctuated between 23.0 (September) to 45.0 ± 0.3 µmol L⁻¹ (June) (Fig. 2 A). In spring monitoring, Si tended to slightly increase from <40.0 µmol L⁻¹ (Spring S1, S2) to 55.0 µmol L⁻¹ in S4 (Fig. 2 B). In the late summer monitoring, Si followed the variations as SRP, with strong depletion in S2 and S3 (Fig. 2 C).

According to Redfield ratios, water mass was N-limited if N/P < 10 and Si/N > 1 (blue area) (N corresponding to $NO_3$ + $NO_2$ + $NH_4$ concentrations), P-limited if Si/P > 22 and N/P > 22 (green area) and Si-limited if Si/P < 10 and Si/N < 1 (pink area) (Justic et al., 1995). During the monthly monitoring, water was first P-limited (February, March and April) and then N-limited (June, August and September) (Fig. 3 A). It was noticed the non-limitation in May, July and October. During the spring monitoring, N limitation was identified in S2, S3 and S4 (Fig. 3 B). In the late summer monitoring, S1 and S4 were N-limited and P-limited in S2 (Fig. 3 C).

#### 3.1.3. Phytoplankton dynamics

Chla was dominated by nanophytoplankton and to a lesser extent by picophytoplankton (Fig. 4). In the monthly follow up, the minimum and
<figure><img src="image_6.png" /><figcaption>**Fig. 1.** Water column temperature, salinity/conductivity, turbidity, and dissolved oxygen (DO concentration and saturation) at L'Houmeau monitored continuously from March 2021 to October 2021. Dotted vertical lines correspond to the field campaigns. Gaps in the data series correspond to faulty equipment. Black arrows represent the draining/replenishment periods of the pond.</figcaption></figure>

<caption>**Table 1**</caption>

Discrete samples of salinity, pH and temperature at L'Houmeau during monthly sample (between February 2021 and October 2021). S1→S4 correspond to weekly sample in spring (May–June) and late summer (September).

<table><thead><tr><th></th><th></th><th>Salinitée(‰)</th><th>pH</th><th>Tempe(°C)</th></tr></thead><tbody><tr><td>Feb</td><td></td><td>26.0</td><td>8.3</td><td>6.6</td></tr><tr><td>Mar</td><td></td><td>28.0</td><td>8.2</td><td>9.7</td></tr><tr><td>Apr</td><td></td><td>32.3</td><td>8.1</td><td>10.1</td></tr><tr><td rowspan="4">May</td><td>S1</td><td>33.9</td><td>8.0</td><td>14.6</td></tr><tr><td>S2</td><td>34.1</td><td>8.0</td><td>15.8</td></tr><tr><td>S3</td><td>35.8</td><td>8.0</td><td>21.6</td></tr><tr><td>S4</td><td>35.5</td><td>8.2</td><td>22.5</td></tr><tr><td>Jun</td><td></td><td>35.7</td><td>8.1</td><td>25.1</td></tr><tr><td>Jul</td><td></td><td>32.2</td><td>8.1</td><td>19.4</td></tr><tr><td>Aug</td><td></td><td>33.4</td><td>8.2</td><td>20.0</td></tr><tr><td rowspan="4">Sep</td><td>S1</td><td>38.4</td><td>8.4</td><td>22.1</td></tr><tr><td>S2</td><td>39.1</td><td>8.6</td><td>21.3</td></tr><tr><td>S3</td><td>35.9</td><td>8.7</td><td>16.6</td></tr><tr><td>S4</td><td>39.9</td><td>8.7</td><td>17.2</td></tr><tr><td>Oct</td><td></td><td>27.8</td><td>8.1</td><td>15.8</td></tr></tbody></table>

maximum biomass were respectively observed in February (<1 µg L⁻¹ for each size class) and June (8.0 ± 1.0 and 5.0 ± 0.5 µg L⁻¹ for nano and picophytoplankton respectively) (Fig. 4 A). During the weekly spring monitoring, nanophytoplankton biomass fluctuated between <2 µg L⁻¹ in S2 to 4.0 ± 0.2 µg L⁻¹ in S3 (Fig. 4 B). In September, the biomass of nanophytoplankton was <8.0 µg L⁻¹ in S1 and S4, but peaked at 27.0 ± 5.0 µg L⁻¹ in Sep S2 (Fig. 4 C).

As for Chla, carbon primary production was largely dominated by nanophytoplankton and to a lesser extent by picophytoplankton (microphytoplankton carbon production was <15 mg C m⁻³ d⁻¹ for all type of monitoring) (Fig. 5). During the monthly monitoring, the average production of these two size classes was <90 ± 10 mg C m⁻³ d⁻¹ between February and August (Fig. 5 A). In September, it increased to 315 ± 35 and about 180 ± 50 mg C m⁻³ d⁻¹ for the nano and picophytoplankton respectively. Production dropped in October for all three size classes (about 100 and 25 mg C m⁻³ d⁻¹ for nano and picophytoplankton respectively). During the spring monitoring (Fig. 5 B), carbon production fluctuated between 38 ± 5 (Spring S2) to 71 ± 28 mg m⁻³ d⁻¹ (Spring S3) for the nanophytoplankton, and between 12 ± 7 (Spring S2) and 38 ± 2 mg C m⁻³ d⁻¹ (Spring S3) for the picophytoplankton (Fig. 5 B). In the late summer monitoring (Fig. 5 C), production peaked

in S3 at 1200 ± 100 and about 700 ± 200 mg C mg m⁻³ d⁻¹ for nano and picophytoplankton respectively, before drop to around 300 mg C mg m⁻³ d⁻¹ for nano and picophytoplankton in S4.

Noted that the maximum Chla concentration and carbon production (in June and September respectively in the monthly follow-ups and globally in the third week of the weekly monitoring) coincided with strong depletion of NO₃⁻, ΣNH₃ and SRP at respective time.

### 3.1.4. Mesozooplankton dynamics

The minimum abundance was observed in February with about 100 ind m⁻³ and the maximum was observed in June with 1150 ind m⁻³. Copepods dominated the mesozooplankton, reaching >95 % of community in September (Fig. 6 A).

During the spring weekly monitoring, the mesozooplankton increased from S1 to reach a maximum about 3500 ind m⁻³ in S3 with nearly 100 % of copepods (Fig. 6 B). In September, the abundance exploded in S2, of the order of 10⁵ ind m⁻³ before drop in S3 to <1000 ind m⁻³ in Sep S4 (Fig. 6 C).

## 3.2. Benthic compartment

### 3.2.1. Organic matter content and phytoplankton

All along the sediment sampling periods, OM fluctuated between 15 and 16 %, slightly decreasing with depth (Fig. 7 A). Chla decreased with depth with surface values varying between 19 and 38 µg g⁻¹ in May and October respectively (Fig. 7 B 6, 12). Note that February, along with October, has the highest Chla concentration at the sediment surface (Fig. 7 B 2, 12). Active Chla (Chlaₐct) generally followed the biomass curves and varied between 42 and almost 60 % at the sediment surface in May and October respectively (Fig. 7 B 6, 12). Overall, the concentration (and active biomass) gradient of Chla was greatest between the surface and 2 cm depth. At the bottom of the core (6 cm depth), Chla biomass was <11 µg g⁻¹ and Chlaₐct ≤ 40 % (Fig. 7 B).

### 3.2.2. Reactive species in the sediment

Profiles of NO₃⁻, NO₂⁻, manganese (dissolved Mnₐ and particulate Mnₐsc), iron (dissolved Feₐ and particulate Feₐsc), phosphorus (SRP and particulate Pₐsc), SO₄²⁻ (Sₐ), ΣNH₃, Si, and alkalinity (Alk) are presented in Fig. 8. Between February and May, NO₃⁻ and NO₂⁻ concentration presented slight decrease from the top to the 2 first cm depth (maximum
<figure><img src="image_7.png" /><figcaption>**Fig. 2.** Surface water concentrations of NO₃⁻, NO₂⁻, ΣNH₃, SRP and Si (mean ± sd) at L'Houmeau during (A) monthly sample (between February 2021 and October 2021), (B) spring and (C) late summer weekly samplings. Vertical scales differ on each follow up.</figcaption></figure>

<figure><img src="image_8.png" /><figcaption>**Fig. 3.** Potential nutrient limitation following the Redfield ratios at L'Houmeau during (A) monthly sampling (between February 2021 and October 2021), (B) spring and (C) late summer weekly samplings. The black lines represent the limitation thresholds of ratio pairs. The stations in the colored zones are N-limited (blue), P-limited (green) and Si-limited (pink).</figcaption></figure>
<figure><img src="image_9.png" /><figcaption>**Fig. 4.** Chlorophyll a (Chla) biomass dynamics (mean ± sd) by size class (micro: microphytoplankton (> 20 µm), nano: nanophytoplankton (3–20 µm), pico: picophytoplankton (< 3 µm)) at L'Houmeau during (A) monthly sampling (between February 2021 and October 2021), (B) spring and (C) late summer weekly samplings. Vertical scales differ on each follow up.</figcaption></figure>

<figure><img src="image_10.png" /><figcaption>**Fig. 5.** Primary production (Prod) dynamics (mean ± sd) by size class (micro: microphytoplankton (> 20 µm), nano: nanophytoplankton (3–20 µm), pico: picophytoplankton (< 3 µm)) at L'Houmeau during (A) monthly sampling (between February 2021 and October 2021), (B) spring and (C) late summer weekly samplings. Vertical scales differ on each follow up.</figcaption></figure>

concentration in February with 0.8 and 2.4 µmol L⁻¹ respectively at the top of the core, Fig. 8. 1). NO₃⁻ globally re-increased from 3 cm depth, to reach 3 µmol L⁻¹ in May at 5 cm depth (Fig. 8. 13). NO₂⁻ seemed to slightly re-increased from 4 cm depth between February and May and from 2 (or even less) cm depth between July and October (up to 0.5 µmol L⁻¹ at 5 cm depth in August).

Mn₉ presented a strong positive gradient from the top (around 0 µmol L⁻¹) to 4 cm depth (240 µmol L⁻¹ in May) between February and May before being >170 µmol L⁻¹ from July all along the core. Conversely, Mnₐₛₙ presented strong decreasing gradients from the top to 2 cm depth (between 12 and 18 mmol kg⁻¹ in May and February respectively) and reached <5 mmol kg⁻¹ at the bottom of the core.

Fe₉ increased with depth, with lowest concentration between February (<10 µmol L⁻¹, Fig. 8. 3)) and May (reached 100 µmol L⁻¹ at 6 cm depth, Fig. 8. 15)). It globally increased from July to October, up to 260 µmol L⁻¹ at 3.5 cm depth (Fig. 8. 33).

SRP profiles followed those of Fe₉, with no or weak enrichment between February and May (< 5 µmol L⁻¹ at the bottom of the core). It strongly increased from July, with a maximum concentration in October with 28 µmol L⁻¹ at 4 cm depth (Fig. 8. 34).

The Pₐₛₙ and Feₐₛₙ profile presented a similar and significant decreasing concentration gradient from the top (65–75 mmol kg⁻¹ for Feₐₛₙ and 20–29 mmol kg⁻¹ for Pₐₛₙ) to 2 cm depth in February, August and October. Below, they presented a constant (Pₐₛₙ) or slight decreasing

(Feₐₛₙ) profile to reach <10 and around 50 mmol kg⁻¹ at the bottom of the core for Pₐₛₙ and Feₐₛₙ respectively.

Between February and May, ΣNH₃ increased from the top (<25 µmol L⁻¹) to the bottom of the core (up to 135 µmol L⁻¹ in May). From July, ΣNH₃ was <4 µmol L⁻¹ between the surface and 4 cm depth and re-increased up to 50 µmol L⁻¹ at 6 cm depth in July (Fig. 8. 23).

Si₉ increased with depth, from 20 µmol L⁻¹ at the surface (February, Fig. 8. 5) to about 150 µmol L⁻¹ at 6 cm (July, Fig. 8. 23). Noted a slight consumption zone in the first cm depth in August and October.

S₉ showed relatively vertical gradients between 20 and 27 mmol L⁻¹ in February and August respectively (Fig. 8. 6 and 30). A slight increase can be noticed in October between the surface and 3 cm (increase of about 2 mmol L⁻¹) (Fig. 8. 36).

Finally, alkalinity was quite constant along the sediment profiles around 5.5 mm kg⁻¹ between February and May. From July it globally increased to a maximum of 7.5 mm kg⁻¹ at 2 cm depth in August (Fig. 8. 30) and up to 11 mmol kg⁻¹ between 3 and 6 cm depth in October (Fig. 8. 36).

### 3.2.3. Nutrient diffusive fluxes at the sediment-water interface

NO₃⁻ globally entered the sediment with maximum influxes in February, April and October with 4.0 ± 5 and 5.5 ± 2.5 and 8.0 ± 4.0 µmol m⁻² d⁻¹ (Fig. 9 A). NO₂⁻ entered the sediment between February and May, from 6.0 ± 2.5 to 0.5 µmol m⁻² d⁻¹ in February and April
<figure><img src="image_11.png" /><figcaption>**Fig. 6.** Mesozooplankton abundances for 9 groups at L'Houmeau during (A) monthly sampling (between February 2021 and October 2021), (B) spring and (C) late summer weekly samplings. Vertical scales differ on each follow up.</figcaption></figure>

respectively. From July to August, $NO_2^-$ presented effluxes around -0.7 µmol m⁻² d⁻¹. (Fig. 9 B).

$\Sigma NH_3$ effluxes were observed between February and July, with maximum in May at -60.0 ± 35.0 µmol m⁻² d⁻¹. It slightly entered the sediment in August, around 5.0 ± 2.5 µmol m⁻² d⁻¹ (Fig. 9 C).

SRP fluxes were almost zero between February and May. It left the sediment in July and October at -7.0 ± 6.5 and -23.0 ± 12.0 µmol m⁻² d⁻¹ respectively and entered-in in August (8.0 ± 2.0 µmol m⁻² d⁻¹) (Fig. 9 D).

Si effluxes seemed to decrease between February and July with -180.0 ± 20.0 and -45.0 ± 25.0 µmol m⁻² d⁻¹ respectively. It entered the sediment in August and October with 20.0 ± 30.0 and around 8.0 ± 16.0 µmol m⁻² d⁻¹ respectively (Fig. 9 E).

# 4. Discussion

## 4.1. Temporality in nutrient limitation in the water column

### 4.1.1. Winter-spring transition: P-limitation

If nitrogen limitation in marine or coastal environments is usually considered (Thomas, 1970; Ryther and Dunstan, 1971; Howarth and Marino, 2006), the present high $NO_3^-$ concentration (almost 40 µmol L⁻¹

in winter and early spring) made the first third of the monitoring (from February to April) P-limited (Fig. 3 A). Such limitation could be shown in freshwater marshes located upstream of this study area (Masclaux et al., 2014; Moncelon et al., 2021, 2022). This high nitrogen concentration could be representative of temperate coastal environments (De Wit et al., 2005; Guillaud et al., 2008; Pineau et al., 2021). The increase of mixing of coastal waters in winter generating a homogenization of the distribution of nutrients in the water column (Morais et al., 2003) or/and the intense nitrification activity could explain such $NO_3^-$ concentration. Moreover, the proximity of supply waters influenced by fluvial inputs in anthropized areas also suggested anthropogenic impact and nitrate discharge from inland agricultural activities (Howarth et al., 2002; Garnier et al., 2021; McLaughlin et al., 2021). Indeed, the study area was located downstream of a watershed dominated by cereal production, where N inputs can be several hundred kg N ha⁻¹ (SAGE Sèvre Niortaise and Charente). The concentrations reached in this study (45 µmol L⁻¹) were almost twice as high as the maxima recorded in coastal areas under fluvial influence slightly further north in the Pertuis Breton (Lannuzel and Polsenaere, 2018).

Numerous works have shown that increased flows, by leaching the catchments, strongly control the different chemical species in the associated water bodies, especially in winter (Guillaud et al., 2008). In the
<figure><img src="image_12.png" /><figcaption>**Fig. 7.** Sedimentary organic matter (OM), chlorophyll-a (Chla) and active Chla (Chla<sub>act</sub>) content profiles (mean ± sd) at L'Houmeau in February, April, May, July, August and October 2021. Values are calculated from dry weight sediment (dw).</figcaption></figure>

context of L'Houmeau pond, this raised the question of diffuse pollution and the importance of inputs from rivers and channels that flow into the Pertuis (Charentais and Breton), such as the Lay, the Sèvre or the Canal du Curé (Polsenaere et al., 2018) especially in winter when rainfall event occurred more frequenlty. Moreover, same trends in the DIN and SRP fluctuations (i. e. synchronous increases or decreases) could testify of regeneration processes (Van Beusekom and De Jonge, 2002) coupling

with photosynthesis processes (Harrison, 1978) in the coastal waters.

### 4.1.2. Spring-summer-fall period: N-limitation

As for winter-spring transition, the fluctuations between DIN and SRP globally followed the same trends between June and October, with strong removal in September. However, the SRP concentration reached slightly a higher level in this second half of the year, while DIN concentrations (especially NO₃⁻) were lower (Fig. 2 A) compared to the winter-spring transition. This places the system in N-limitation, particularly in June, August and September (Fig. 3 A). The NO₃⁻ deficit could be caused by both limited allochthonous input and pumping by primary producers. The first hypothesis could be justified by the fact that the lowest average NO₃⁻ concentrations (<13 µmol L⁻¹) were observed when salinity levels were highest, which could indicate a lower input by drainage of NO₃⁻ loaded continental waters. Pumping could be justified by the presence of more abundant phytoplankton communities from June (Fig. 4 A), although the carbon primary production was maximal in September (Fig. 5 A). The relative increase in SRP concentrations (compared to DIN) suggested the contribution of an autochthonous source from the pond, as the benthic compartment, that mobilize P in the water column (from July to October, Fig. 8. 21, 27, 33), and generating a shift in nutrient limitation.

### 4.1.3. Influence of the sediment for water column nutrient limitation

The sediment contribution for water P inputs has been suggested in numerous works in fresh (Caraco et al., 1990; Lamers et al., 2002; Malecki et al., 2004; Herzsprung et al., 2010; van Dael et al., 2020), saline (Caraco et al., 1990; Eilola et al., 2014), or transitional environments (Hartzell and Jordan, 2012). Here, shift from P to N limitation coincided with period of SRP mobilization in the porewater during spring-summer and summer-fall transition, where SRP effluxes were observed in July and October. However, SRP mobilization (and N limitation in the water column) did not always coincide with SRP effluxes from the sediment, especially in August where influx was observed (Fig. 9 D). This could be due to an SRP depletion at the SWI in link with microphytobenthos (MPB) development at that time (Fig. 7 B 10) (Dunn et al., 2013). Its presence was confirmed with the Si<sub>d</sub> consumption zone in the first 2 cm depth of sediment (Fig. 8. 29).

In addition, it has been shown that molecular diffusion alone could not be the dominant process controlling nutrient exchanges (Mu et al., 2017). Bioirrigation, bioturbation or advection can occur and increase the real nutrient exchanges at the SWI (Haese, 2005). This was supported since N limitation stage remained stable in August despite non diffusive SRP effluxes. Note that despite the increase of SRP effluxes in October, the system was not nutrient limited. This was explained by the re-increase of DIN compounds, probably due to a quick N recycling in coastal environment after summer bio-productive period.

## 4.2. Nutrient mobilization in the sediment

### 4.2.1. Winter-spring period

Between February and May, no significant SRP mobilization was observed in the first 4 cm of sediment. Despite the low NO₃⁻ concentrations in the water column compared to the freshwater marshes (Moncelon et al., 2021, 2022), slight depletion of nitrate and nitrite in the top two cm of sediment in February could suggest the presence of denitrification processes (Fig. 8.1). It can be noted that the NO₃⁻ increase from 2 cm depth from February to May (Fig. 8. 1, 7, 13) might be due to more in depth nitrification or anaerobic ammonium oxidation (anammox) processes since it coincided with ΣNH₃ increase and effluxes at that time. Increase in Mn<sub>d</sub> concentration and the consumption of Mn<sub>asc</sub> in the first 4 cm of sediment indicated the respiration of manganese oxides from the start of the monitoring (Fig. 8. 2). In April and May, Mn<sub>d</sub> mobilization seemed to intensify and reach the SWI, while denitrification processes decreased.

Below 4 cm depth, a slight increase of SRP was observed, coinciding
<figure><img src="image_13.png" /><figcaption>**Fig. 8.** Dissolved nitrates ($NO_3^-)$, nitrites ($NO_2^−$), manganese ($Mn_d$), iron ($Fe_d$), soluble reactive phosphorus (SRP), sulphates ($S_d$), ammonium ($\Sigma NH_3$), silica (Si), and alkalinity (Alk) (empty circles), reductively extracted manganese, iron and phosphorus ($Mn_{asc}$, $Fe_{asc}$, $P_{asc}$) (filled circles) (mean ± sd) along the sediment column from 0 to 6 cm depth at L'Houmeau in February, April, May, July, August and October 2021. The arrows represented the concentration in the water column.</figcaption></figure>

with that of dissolved iron. This could testify of the beginning of iron respiration processes especially in April and May (Fig. 8, 10, 16) that release SRP in porewater (Moncelon et al., 2021).

### 4.2.2. Summer-fall period

The decrease of $NO_3^−$ in the porewater coincided with that of $\Sigma NH_3$ and the global increase of $Mn_d$ between July and October comforted the potential anaerobic nitrification in winter spring transition (Javanaud, 2011).

From July, the pore waters were characterized by the increase of SRP and alkalinity production announcing a shift in the mineralization processes of the OM. It coincided with $Fe_d$ increase in the same layers of sediment (Fig. 8). This testified of a strong coupling between iron oxide respiration for OM mineralization and release of P in porewater through Fe—P desorption, notably in marsh system (Chambers and Odum, 1990; Moncelon et al., 2021). This process can potentially be favoured by the increase of temperatures at that time (Li and Gregory, 1974; Bullock et al., 2013), allowing a diffusive SRP efflux toward the water column (Fig. 9 D). In addition, increasing salinities over the season might also promote P desorption in the sediment (Stumm and Morgan, 1981; Sundareshwar and Morris, 1999; Loomis and Craft, 2010; Zhang and Huang, 2011). Noted that the P mobilization temporality observed in

this work was at odds with work highlighting increased sediment P retention with increasing temperatures (Huang et al., 2011; Zhang and Huang, 2011; Bai et al., 2017).

Moreover, anaerobic ammonium oxidation, coupled to iron (III) reduction could be another process for P mobilization. Indeed, the development of the $Fe_d$ peak corresponded to the collapse of the ammonium gradient (Fig. 8, 21, 23), and SRP efflux (Fig. 9 D) from July to October (Zhang et al., 2024). Noted that anammox coupled to manganese reduction could also exist since increased in $Mn_d$ concentration have been observed at that time compared to the winter-spring period (Cheng et al., 2022).

The upturn in Fe(III) respiration processes between the winter-spring and summer-fall periods was clearly visible in the 2D gels (Fig. S2, supplementary materials), where the mobilization of $Fe_d$ and SRP appeared to rise in the sediment column by about 3 cm between May and August–October. Noted that the re-increase of metal oxide and especially $Fe_{asc}$ from July to October in the 2 first cm (Fig. 8, 21, 27, 33) could limit SRP efflux (no efflux in August) at that time by favouring Fe—P formation. This was supported by the synchronous increase of $P_{asc}$ in the same layers of sediment. These increases could be caused by an additional input from the overlying water column, as suggested by the increase of suspended material from July (Fig. S3, supplementary
<figure><img src="image_14.png" /><figcaption>**Fig. 9.** Average NO₃⁻ (A), NO₂⁻ (B), ∑NH₃ (C), SRP (D) and Siₐ (E) diffusive fluxes (mean ± sd) at the sediment-water interface at L'Houmeau in February, April, May, July, August and October 2021. The negative flux values corresponded to fluxes leaving the sediment, the positive values to fluxes entering the sediment. Vertical scales differ on both sites.</figcaption></figure>

materials). Indeed, it has been shown in some estuarine systems that >90 % of P brought into the system was associated with suspended solids and much of that was bound to Fe(III) (Follmi, 1996). Moreover, the O₂ production by MPB at that time could have led to Fe oxidation (Feasc max in August coincided with increase in MPB), promoting the development of the iron curtain (Delgard et al., 2012; Sundbäck and Granéli, 1988), cancelling the SRP efflux in August (Fig. 9 D). The SRP influxes observed in August despite the strong SRP mobilization in porewaters (Fig. 8. 28) could also be explained by the presence of O₂ caused by bioturbation (Thibault de Chanvalon et al., 2017). Indeed, bioturbation marks have been observed in August in the 2 first cm of sediment (Fe₉ and SRP 2D profile, Fig. S2). These resulted in the local decrease of Fe₉ and SRP concentration and increased that of oxides (Fig. 8. 27, 28), thus cancelling SRP efflux.

Noted that the re-increase of SRP effluxes in October despite the significant presence of MPB could be explained by its potential senescence. The high Chla content in the top 0.5 cm of sediment in October (Fig. 7 B 12) could be due to the sedimentation of the phytoplankton bloom of September, the most productive period (Fig. 5 A). Indeed, the phytoplankton sedimentation is a relatively rapid phenomenon (between 0.18 and 2.96 m/day; Smayda and Boleyn, 1965). The potential resulting reduction of O₂ concentration in the sediment surface layers could thus counteract the formation of Feasc and P trapping capacity. This was supported by the decrease of Feasc at the SWI in October (–25 mmol·kg⁻¹, Fig. 8. 33).

### 4.2.3. Organic matter mineralization efficiency

The high Chla content that reach the sediment compartment in these systems (MPB or phytoplankton sedimentation) constituted a supply of labile OM that can be rapidly operated via anaerobic mineralization processes (Laverman et al., 2021). However, the non-variation of OM profiles (Fig. 7 B) testified of a large OM stock exceeding the substrate demand for mineralizers (Moncelon et al., 2022) or, most likely, a high carbon refractory content in this type of environment. Even though sulphate reduction is widely recognized as one of the major processes in marine and coastal environments for anaerobic mineralization of OM (Canfield et al., 1993), the almost non-existent S₉ gradients in the top 6

cm of sediment suggested that sulfate-reduction processes occurred deeper due to the thickness of the iron curtain supplied by the sedimentation of iron oxides-enriched SPM. Variations in alkalinity, often correlated with the presence of sulfate reduction in a salty environment, seemed here, by its convex side between July and October (Fig. 8. 24, 30, 33), to be more closely linked to metal recycling. The absence of sulfate reduction in the top 6 cm of the sediment could promote P diffusion toward the water column by inhibiting the formation of iron curtain through FeS complex production (Rozan et al., 2002).

## 4.3. Phytoplankton dynamic and drivers

### 4.3.1. Phytoplankton typology

The dominance of small size classes (nano and picophytoplankton) and no real decrease in Si concentration in the water column was observed suggesting the low contribution of diatoms at that time (Glé et al., 2008). This also confirmed a nutrient-limited environment in N or P, where small cells competitiveness is favoured according to their higher specific ratio favouring diffusion processes across the membrane (Masclaux et al., 2014). Moreover, this dominance could be explained by relatively high-water temperatures, especially in spring (means close to 25 °C). Indeed, works on the size-temperature relationship has suggested that smaller organisms are more favoured than larger ones in warmer conditions due to faster metabolic processes (Daufresne et al., 2009; Sommer et al., 2017).

### 4.3.2. Phytoplankton dynamic and nutrient availability

The limitation shifts observed in the monthly follow-up (Fig. 3 A) make it difficult to explain phytoplankton bloom phenomena according to nutrient availability and thus prevent prioritization of potential nutrient management actions (Conley, 2000; Dupuy et al., 2007). If it has been shown the direct link between limiting nutrient effluxes from sediment and phytoplankton development in freshwater marsh systems (Moncelon et al., 2021, 2022), this work highlights more complex interaction between nutrient availability and primary producers dynamic. Phytoplankton development, based on Chla measurements, showed here classical seasonal pattern with bloom in spring and late
summer (Fig. 4). The Chla peak observed in June occurred under N-limited conditions. The almost zero N concentrations at that time highlighted the important pumping exerted by the phytoplankton, the main cause of this limitation. Noted that no nutrient limitation was observed before the bloom, i. e. in May.

The second bloom observed resulted in a peak of carbon production in September, with a much lower Chla peak than that of June (Fig. 4 A, Fig. 5 A). This might highlight stronger predation efficiency pressure at that time (Liu et al., 2019; Sautour et al., 2000; Kouassi et al., 2001). Indeed, grazing by zooplankton represents the major fate of marine phytoplankton biomass (Banse, 2013; Steinberg and Landry, 2017). It could lead to the potential dominance of respiratory processes over primary production (Lajaunie-Salla et al., 2017) as it could be seen by the relatively low O₂ concentration, especially during spring bloom event (Fig. 1, Table 1).

The high carbon production in September coincided with very low nutrient concentrations (~1 µmol L⁻¹) and was preceded by higher SRP concentrations in July and August, Fig. 2 A) suggesting hysteresis (Wirtz and Smith, 2020) or high carbon production rate when the observed nutrient loading is low (de Jonge, 1997). Moreover, the global decrease in NO₃⁻ concentration in the July–October period (compared to Febru ary–May period, Fig. 2 A) could suggest a shift in N sources for primary production. The NO₃⁻ pulses in February and April, before the bloom period, could reflect the use of allochthonous N, while the lower nutrient peaks in July and October, post bloom events, suggested the supply of autochthonous nutrients by regeneration. The overall increase in all nutrients at the end of each weekly monitoring (Fig. 2 B, C) in enclosed systems points toward regeneration (in both water column and sedi ment). The high phytoplankton production accompanied by important nitrate decreas in the water column in the weekly monitoring of late summer compared to June evidenced a rapid turnover and support the idea of a primary production dominated by regenerated by-products (Metzler et al., 1997).

Mesozooplankton influence on nutrient regeneration is widely recognized, especially on N cycling in tropical and temperate coastal environment (Hernández-León et al., 2008). The high phytoplankton production rates in September coincided with the DOC maxima (a DOC peak was also observed in June coinciding with the Chla peak) (Fig. S3 A, supplementary materials) probably caused by the exudation of extracellular material performed by photosynthetic organisms (Marañón et al., 2004). It could also be indicative of the predation pressure on the phytoplankton compartment (sloppy feeding, excretion) since peaks of DOC in June and September coincided with peaks of mesozooplankton (Fig. 6 A). These DOC increases (also observed during weekly monitoring, Fig. S3 B, C, supplementary materials) could support heterotrophic bacterial communities (Dupuy et al., 2011) and support a system driven by nutrient regeneration processes in the water column at that time (Mostajir et al., 2015). Noted that the decrease in meso-zooplankton abundance and DOC concentration in July and August could be attributed to the decrease in temperature near 19 °C at that time. This temperature has recently been shown to be the threshold at which copepod growth is strongly impacted in confined coastal salt marshes (Quintana et al., 2021).

The drop of carbon production in October (Fig. 5 A) seemed to allow the regeneration of all nutrients and the non-limiting conditions (Fig. 2 A). In addition to the drop in the water temperature, the high SPM concentrations (>120 mg L⁻¹, Fig. S4, supplementary materials) at that time could reflect coastal water mixing or wind-related resuspension in the pond, which could have limited planktonic primary production (Borges and Abril, 2011; He et al., 2017; Golubkov and Golubkov, 2022). The increase in general O₂ concentration in October (Fig. 1) could suggest the development of other primary producers such as macroalgae, MPB or even new small cells phytoplankton bloom in the basin.

## 4.4. Assessment of the purifying capacity of the pond in a benthos-pelagos context

### 4.4.1. *Nutrient mitigation: Phytoplankton pumping vs regeneration*

During the weekly spring monitoring, DIN decreased of about 75 % in the pond. Although denitrification can be a major process for nitrogen mitigation in pond environments subject to agricultural pressure (Lv et al., 2024), the geochemical data at that time (i.e. in May and June) seemed to not attest a sufficient contribution of denitrification in the sediment that could justify the decrease of NO₃⁻ in the water column. This illustrated the phytoplankton pumping efficiency, for a total DIN mitigation occurring within days in the pond. Conversely, Si increased slightly (25 %) and P returned to almost its initial levels (increase <1 %) indicating an internal recycling of P (benthic and pelagic) lower or equal to 4 weeks.

In contrast, the late summer weekly monitoring indicated a 3-week water retention time for a maximum nutrient removal and a DIN increase of almost 290 % and a near return to initial conditions for SRP and Si at the end of the monitoring (Fig. 2 C). The nitrogen fluctuation could be controlled by the succession between organic N remineralization, heterotrophic nitrification and NO₃⁻ assimilation promoted by phytoplanktonic blooms (Chen et al., 2021). The increase of SRP could be both attributed to the pelagic regeneration and release of SRP from the sediment. Simultaneously, the phytoplankton decline (biomass and carbon production) in Sep S4 could be attributed to a nutrient limitation, particularly in nitrogen that could follow the important production in Sep S3 and a not fast enough regeneration of nutrients.

Enclosed system monitoring highlighted the importance of the water renewals frequency in clear bed type systems in order to promote the purification of water masses and limit the risks of eutrophication in coastal zones. Shift in nutrient limitation and potential nutrient resupply in the marsh system (e. g. Spring monitoring) could have a chemical and ecological impact on receiving waters. Here, the water from the pond at the end of the Spring monitoring was more DIN rich than that of supply water in June, counteracting its water purification function. In September, despite the increase in nutrient concentration in the enclosed system, the water remained less concentrated in nutrients than the supply water in October, thus limiting the risk of eutrophication of the coast from the water exported from the pond.

### 4.4.2. *Sediment contribution for water column nutrient enrichment*

The re-increase of nutrient at the end of each weekly monitoring did not appear to involve all nutrients in the same way. In spring, the lower observed DIN regeneration compared to SRP in the water column could be caused by simultaneous nitrification and aerobic denitrification processes (Robertson and Kuenen, 1984; Chen et al., 2018). Indeed, the physical-chemical conditions (temperature, pH and O₂ concentration) at the end of monitoring (Fig. 1, Table 1) broadly corresponded to those in which aerobic denitrifiers tend to work efficiently (i.e. 25–37 °C and pH 7–8, DO 3–5 mg L⁻¹, Ji et al., 2015). However, DOC:DIN ratios ranging from 0.5 to 3 appeared to be below the ranges reported to promote an effective aerobic denitrifiers activity (Ji et al., 2015). A more likely explanation would be to consider sediment as a source of SRP.

As for the spring monitoring, the excess of SRP (despite the increase in DIN) over the Redfield ratios at the end of the late summer weekly monitoring suggests additional autochthonous input that could also cause the N-limitation. Based on the October benthic fluxes (outflow of 20 µmol m⁻² d⁻¹) (October campaign was carried out 2 weeks after that of S4), a quick calculation allowed to estimate at 0.3 µmol L⁻¹ the increase in SRP in the water column via the SRP efflux in one week, that was between Sep S3 and Sep S4. This calculation was based on an overlying water volume of 90 m³ (pond surface: 180m² and 0.5 m of water height) and corresponded to near 40 % of the SRP increased between Sep S3 and Sep S4.

Noted that the contribution of nutrient input from the sediment to phytoplankton demand can also be considered. This contribution could
potentially reach 15 and 60 % of phytoplankton P demand during bloom (September) and non-bloom event (October) respectively, and >10 % of DIN demand in non-bloom period (May). Irrelevant sediment contribution for DIN demand during bloom event was recorded since DIN effluxes were too weak at that time. These shifts in contribution between bloom and non-bloom event highlighted the importance of pelagic input (allochtonous or autochtonous) during bloom event. This compares with other coastal environments where the sediment P contribution for primary production was roughly the same during bloom and non-bloom events. This indicated a more marked proportionality between benthic and pelagic compartment interactions over time (Grenz et al., 2000). More generally, it has been shown that benthic nutrient regeneration can provide 7.5–22 % (Mortazavi et al., 2012; Trimmer et al., 1998) of nutrients to sustain primary production in estuarine or open sea systems (Sun et al., 2021) and even over than 100 % in some estuaries (Blackburn and Henriksen, 1983) or freshwater drained marshes (Moncelon et al., 2022).

Moreover, it would appear that nutrient accumulation can still occur in the water column, even if the sedimentary inputs does not exceed the phytoplankton demand (>100 %). In Sept S4, sediment contribution for SRP water supply reached 40 % whereas the sediment efflux represented only 12 % of phytoplankton demand. Indeed, nutrient regeneration processes in the water column can make a major contribution to phytoplankton primary production in enclosed bays (La Roche, 1983), lakes (Axler et al., 1981; Dodds et al., 1991; Sterner et al., 1995; Vanni, 2002) or marine environments (Harrison, 1980; Clark et al., 2016). In lake environments, P regeneration via zooplankton excretions could contribute between 4 % and 58 % of phytoplankton demand (Vanni, 2002). In enclosed bays, regenerated ammonium can contribute to 36 % of phytoplankton primary N production (La Roche, 1983) and up to 80 % in oceanic environments (Harrison, 1980). In this work, the nearly non input of DIN from the sediment in summer-fall period (especially $\sum\text{NH}_3$) highlighted the importance of pelagic N regeneration for phytoplanktonic production.

## 5. Conclusion and perspectives

Biogeochemical factors governing nutrient dynamics in the water column were explored in a salt marsh system of a clear bed type. A combined monthly and weekly follow up in open and enclosed system evidenced the role of the sediment in changes of nutrient limitation of the water column. This was especially enhanced by SRP effluxes during the spring-summer transition caused by the mineralization of organic matter through iron oxide reduction and anammox coupled to iron (III) reduction processes in the sediment.

The purification capacity of the experimental pond strongly depended on the water retention time, intensity of phytoplanktonic bloom and pelagic or benthic nutrient regeneration. While the latter did not drive nutrient mitigation, it did contribute to the enrichment of the water in SRP and may even have favoured the development of phytoplankton. This raises water management issues for nutrient exportation to the coast (in dissolved or particulate forms), which is a major challenge for nutrient regulation within the land-sea continuum and limiting the eutrophication risk in coastal areas. While the eutrophication potential associated with nitrogen seemed to be reduced in such pond, that of phosphorus needs to be more considered. Actionable mitigation measures such as sediment dredging, phytoremediation or regular water discharge could limit that risk. Moreover, the single-year dataset (2021) may not capture interannual variability or extreme climate impacts. Long-term monitoring is needed to cover consequences of such event and better assess purification capacity of ponds.

### CRediT authorship contribution statement

**Raphaël Moncelon:** Writing – review & editing, Writing – original draft, Visualization, Methodology, Investigation, Formal analysis, Data

curation, Conceptualization. **Charlotte de Lignières:** Visualization, Validation, Methodology, Investigation, Formal analysis. **Philippe Pineau:** Validation, Formal analysis. **Claire Emery:** Validation, Formal analysis. **Eric Bénéteau:** Validation, Formal analysis. **Edouard Metzger:** Validation, Supervision, Methodology, Investigation, Formal analysis, Conceptualization. **Christine Dupuy:** Validation, Supervision, Methodology, Investigation, Formal analysis, Conceptualization.

### Declaration of competing interest

The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper.

### Acknowledgments

This study was supported by the « Ministère de l'Enseignement Supérieur et de la Recherche », the water agencies Loire-Bretagne and Adour-Garonne, the General Council of Charente-Maritime, the European Union. This research was supported by a PhD grant from the La Rochelle conurbation (CDA). We acknowledge the LIENSs AutoAnalyzer platform for nutrient analyses and the "Radioecology lab" for production measurements and the LPG geochemical analytic platform for ICPAES analyses. Thanks are extended to the UNIMA (Union des Marais de Charente-Maritime), FMA and stakeholders for their expertise on the Charente-Maritime marshes and to LRTZC (La Rochelle Territoire Zéro Carbone) for the post-doctoral wage.

## Appendix A. Supplementary data

Supplementary data to this article can be found online at https://doi.org/10.1016/j.scitotenv.2025.179308.

### Data availability

Data will be made available on request.

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