Efficient removal of phosphorus and nitrogen in sediments of the eutrophic Stockholm Archipelago, Baltic Sea

Coastal systems can act as filters for anthropogenic nutrient input into marine environments. Here, we assess the processes controlling the removal of phosphorus (P) and nitrogen (N) for four sites in the eutrophic Stockholm Archipelago. Bottom water concentrations of oxygen and P are inversely correlated. This is attributed to the seasonal release of P from iron (Fe)-oxide-bound P in surface sediments and from degrading 20 organic matter. The abundant presence of sulfide in the pore water, linked to prior deposition of organic-rich sediments in a low oxygen setting (“legacy of hypoxia”), hinders the formation of a larger Fe-oxide-bound P pool in winter. Burial rates of P are high at all sites (0.03-0.3 mol m -2 y -1 ), a combined result of high sedimentation rates (0.5 to 3.5 cm yr -1 ) and high sedimentary P at depth (~30 to 50 μmol g -1 ). Organic P accounts for 30-50% of reactive P burial. Apart from one site in the inner archipelago, where a vivianite-type 25 Fe(II)-P mineral is likely present at depth, there is little evidence for sink-switching of organic or Fe-oxide bound P to authigenic P minerals. Denitrification is the major benthic nitrate-reducing process at all sites (0.09 to 1.7 mmol m -2 d -1 ), efficiently removing N as N2. Denitrification rates decrease seaward following the decline in bottom water nitrate and sediment organic carbon. Our results explain how sediments in this eutrophic coastal system can efficiently remove land-derived P and N, regardless of whether the bottom waters are oxic or 30 frequently hypoxic. Hence, management strategies involving artificial reoxygenation are not expected to be successful in removing P and N, emphasizing a need for a focus on nutrient load reductions.

particularly in low-salinity environments with high inputs of Fe-oxides (e.g. Egger et al., 2015). Burial of P is redox sensitive, with retention of P bound to Fe-oxides and in organic matter decreasing upon increased hypoxia and anoxia (e.g. Van Cappellen and Ingall, 1994). However, a more limited exposure to O2 also enhances the preservation of organic matter, and may allow organic P to become the dominant form of P in the sediment (Lukkari et al., 2009;Mort et al., 2010;Slomp, 2011). between N removal (as N2) from the ecosystem or transformation of organic-N to NH4 + , which can be retained in 80 the ecosystem, may be strongly influenced by eutrophic conditions.
Predictions of the response of coastal areas to decreased nutrient inputs and/or natural or artificial reoxygenation require insight in the processes responsible for P and N cycling and whether P and N are transformed, retained or removed. This is of particular relevance to the coastal zone of the Baltic Sea because of its highly eutrophic and frequently low O2 state (Conley et al., 2011). Active nutrient reductions from the 1980s onward (Gustafsson et 85 al., 2012) are now leading to the first signs of recovery in the region (Andersen et al., 2017). A good example of a recovering system within the Baltic Sea is the Stockholm Archipelago, where recovery from hypoxia (Karlsson et al., 2010) may be associated with increased P burial (Norkko et al., 2012). Based on coupled physical and biogeochemical models it was recently suggested that the Stockholm Archipelago was very efficient in removing P and N for the period 1990-2012, accounting for loss of 65 % of the land-derived P input and 75 % of the land-90 derived and atmospheric N input (Almroth-Rosell et al., 2016). The area-specific P and N retention was highest in the inner part of the Stockholm Archipelago. Based on the high NO3concentrations in the bottom water, and high organic carbon contents in the sediment in the Archipelago, benthic denitrification is expected to dominate N removal (Almroth-Rosell et al., 2016;Asmala et al., 2017). Recent mass balance modelling for the inner Archipelago suggests that sediments are a P sink in winter and a source in summer and autumn, with low annual 95 net retention in the sediments (Walve et al., 2018). These apparently conflicting results between different modelling approaches emphasizes the need to better understand and quantify P removal, i.e. permanent burial of P in the sediment.
The objectives of this study are to identify and quantify the main P-burial phases and the processes controlling removal of N in sediments of the Stockholm Archipelago and to determine the time scales that govern removal 100 and the implications for management strategies. We present geochemical depth profiles for a range of sediment components (P, Fe, organic carbon) and rate measurements of benthic N cycling processes for four sites along a gradient from the inner archipelago towards the open Baltic Sea. These sites capture a range of bottom water O2 concentrations from seasonally hypoxic/occasionally euxinic to oxic. Our results highlight the key processes in sediments in eutrophic coastal systems that lead to removal of P and N and that may prevent their further transport due to the differences in salinity between the (nearly) fresh surface water and the underlying more saline water.
In the summer, water column stratification is more pronounced and widespread due to the development of a 120 thermocline. However, in the more open parts of the archipelago, wind-driven mixing may interrupt stratification (Gidhagen, 1987).
The average annual nutrient input into the Stockholm Archipelago was 217 t P and 8288 t N for the period 1990-2012, of which approximately 174 t P and 5846 t N entered the inner archipelago via the Norrström river (Almroth-Rosell et al., 2016). This high nutrient load mostly originates from wastewater treatment facilities of Stockholm 125 (Johansson and Wallström, 2001) and, in combination with (seasonal) stratification of the water column, led to widespread eutrophication in the past. As a result, large parts of the Stockholm Archipelago are or have been (seasonally) hypoxic to euxinic over the past century (Jonsson et al., 1990;Conley et al., 2011). Studies have shown decreases in dissolved inorganic P and total P due to reductions in nutrient inputs from sewage treatment plants (Walve et al., 2018) and indications of environmental recovery have been deduced from visual observations 130 of sediment cores (Karlsson et al., 2010).  Table 1; Sup. Fig. 1). Extensive water quality monitoring of the study area by the Swedish Meteorological and Hydrological Institute (SMHI, 2019), shows a clear inverse correlation between bottom water O2 concentrations and P and a positive correlation between bottom water O2 and N/P-ratios ( Fig. 3a, b). Bottom water O2 and nutrient concentrations follow a distinct annual pattern, with maximum O2 and minimum nutrient concentrations in winter. After winter, O2 gradually drops and nutrient For anoxic sediment and pore water collection, one core was sliced in a N2-filled glove bag. Two bottom water samples were taken from the overlying water after which the core was sliced at a resolution of 0.5 cm (0 to 10 cm), 2 cm (10 to 20 cm), 4 cm (20 to 40 cm) and 5 cm until the bottom of the core. The sediment was centrifuged (in 50 mL tubes) at 3500 rpm for 20 minutes to extract pore water. The sediment remaining after centrifugation was stored in N2-flushed gas-tight aluminum bags at -20 °C until further analysis. Bottom and pore water samples were filtered over a 0.45 µm filter in a N2-filled glove bag. Subsamples were taken for (1) H2S analysis (0.5 mL was added to 2 mL 2 % zinc (Zn)-acetate); (2) analysis of dissolved Fe and P (1 mL was acidified with 10 μL 30 % suprapur HCl); (3) analysis of sulfate (SO4 2-) (0.5 mL), and stored at 4°C. Subsamples for N-oxides (NOx = NO3 -+ nitrite (NO2 -); 1 mL) and NH4 + (1 mL) were stored at -20 °C.

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At Strömmen, one core was sliced at the same resolution as described above to determine porosity and 210 Pb. Data for porosity and 210 Pb for the other three study sites were taken from van Helmond et al. (in review).

Bottom and pore water analysis
Concentrations of CH4 were determined with a Thermo Finnigan Trace gas chromatograph equipped with a flame ionization detector as described by Lenstra et al. (2018). The average analytical uncertainty based on duplicates 170 and triplicates was <5 %. Pore water H2S was determined spectrophotometrically using phenylenediamine and ferric chloride (Cline, 1969). Upward fluxes of H2S in the porewater towards the sediment surface were calculated as detailed in Hermans et al. (2019a). Dissolved Fe and P (assumed to be present as Fe 2+ and HPO4 2-) were measured by Inductively Coupled Plasma-Optimal Emission Spectroscopy (ICP-OES; SPECTRO ARCOS).
Concentrations of SO4 2were determined by ion chromatography. The average analytical uncertainty based on duplicates was <1 %.

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All sediment samples were freeze-dried, powdered and homogenized using an agate mortar and pestle in an argonfilled glovebox. Prior to analysis, samples were split into oxic and anoxic fractions (i.e. samples stored open to air and in a N2 or argon atmosphere).

Total elemental composition
Approximately 125 mg of the oxic sediment split was digested in a mixture of strong acids as described by van 185 Helmond et al. (2018). The residues were dissolved in 1 M HNO3 and analysed for their elemental composition by ICP-OES. Average analytical uncertainty based on duplicates and triplicates was <5 % for calcium (Ca) and <3 % for P. The calcium carbonate content (CaCO3 wt.%) was calculated based on the Ca content measured by ICP-OES, assuming that all Ca was in the form of CaCO3.

Organic carbon and nitrogen
a Fisons Instruments NA 1500 NCS analyzer. Average analytical uncertainty based on duplicates was <2 % for carbon and <3 % for nitrogen. Organic carbon (Corg) and nitrogen (Norg) contents were calculated after a correction for the weight loss upon decalcification and the salt content of the freeze-dried sediment. For Baggensfjärden, Erstaviken and Ingaröfjarden C and N contents were taken from van Helmond et al. (in review).

Sequential extraction of iron
Between 50 and 100 mg of the anoxic sediment split was subjected to a sequential extraction procedure based on a combination of the procedures by Poulton and Canfield (2005) and Claff et al. (2010) to determine the different phases of sedimentary Fe (Kraal et al., 2017). Briefly, under O2-free conditions: (1) 10 mL 1 M HCl, pH 0 was 200 added to extract (4 h) Fe(II) and Fe(III) minerals such as easily reducible Fe-oxides (e.g. ferrihydrite and lepidocrocite), Fe-carbonates and Fe-monosulfides; (2) 10 mL 0.35 M acetic acid/0.2 M Na3-citrate/50 g L -1 Na dithionite, pH 4.8 was added to extract (4 h) crystalline Fe oxide minerals such as goethite and hematite; (3) 10 mL 0.17 M ammonium oxalate/0.2 M oxalic acid, pH 3.2 was added to extract (6 h) recalcitrant oxide minerals such as magnetite; (4) 10 mL 65 % HNO3 was added to extract (2 h) pyrite (FeS2). For all extracts, Fe 205 concentrations were determined colorimetrically with the phenanthroline method, adding hydroxylaminehydrochloride as a reducing agent to convert all Fe 3+ into Fe 2+ (APHA, 2005). For the first step the absorbance before and after addition of the reducing agent was measured, in order to separate Fe 2+ and Fe 3+ . The Fe concentrations of the Fe 3+ fraction of the first step and the second step were summed, and are henceforth referred to as Fe-oxides. Average analytical uncertainty based on duplicates and triplicates was <10 % for all fractions.

Sequential extraction of sulfur
Approximately 300 mg of the anoxic sediment split was subjected to a sequential extraction procedure (Burton et al., 2008) to determine sedimentary sulfur phases. Briefly, under O2-free conditions: (1) 10 mL 6 M HCl and 2 mL 0.1 M ascorbic acid were added to dissolve acid-volatile sulfur (AVS, assumed to represent Fe-monosulfides -FeS) and the released H2S was trapped in a tube filled with 7 mL alkaline zinc acetate solution (24 h); (2) 10 mL 215 acidic chromium(II)chloride was added to dissolve chromium-reducible sulfur (CRS, assumed to represent FeS2) and the released H2S was trapped in a tube filled with 7 mL alkaline zinc acetate solution (48 h). For both fractions, the amount of sulfur in the zinc sulfide precipitates was determined by iodometric titration (APHA, 2005).
Average analytical uncertainty, based on duplicates, was <7 % for both AVS and CRS.

Sequential extraction of phosphorus
organic matter (Org. P). The P content in the citrate-dithionite-bicarbonate extract was analysed by ICP-OES. All other solutions were measured colorimetrically (Strickland and Parsons, 1972). Average analytical uncertainty, based on duplicates, was <7 % for all fractions. Total P derived from acid digestion and subsequent ICP-OES analyses was on average within 5 % of the summed P fractions derived from the sequential extraction.

15 N incubations
Rates of benthic NO3 --reducing pathways were determined using the whole-core isotope pairing technique (IPT) and parallel slurry incubations (Nielsen, 1992;Risgaard-Petersen et al., 2003). Bottom water from Niskin bottles collected at each site was used to fill the incubation chamber (approx. 30 L) and maintained at in situ O2 concentrations using compressed air and nitrogen gas mixtures. Small core liners (Ø 2.5 cm) were used to take 240 sub-cores from the Gemini cores and were immediately transferred to the incubation tank so that all cores were submerged and stoppers were removed. Sodium 15 N-nitrate solution (Na 15 NO3, 98 atom % 15 N, Sigma Aldrich, final concentration ~50 µmol L -1 ) was added to the water of the incubation tank and cores were pre-incubated in the dark at in situ temperature for 2 to 5 h. Three replicate cores were sacrificed by slurrying the entire sediment volume at approximately 0, 2, 5 and 8 h following pre-incubation. Sediment was allowed to settle for 2 minutes 245 before samples for gas (12 mL exetainers, Labco, UK, killed with 250 µL zinc chloride solution, 50 % w/v) and nutrients (10 mL, killed with 250 µL zinc chloride solution, frozen) were taken.
Sediment slurries were carried out in parallel to whole-core incubations. Briefly, a glass bead (0.5 cm Ø) was added to each 12 mL exetainer, which was then filled with filtered (0.2 µm) helium-purged bottom water.
Homogenised surface sediment (2 mL, 0-2 cm depth horizon) was added to each exetainer and vials were sealed.

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Exetainers were incubated on a shaker table in the dark at in situ temperature for 8 to 12 h ensuring consumption of background NO3and O2 before addition of 15 N-substrates. Exetainers were divided into two treatments, amended with sodium 15 N-nitrate or with sodium 14 N-nitrite and 15 N-ammonium chloride (each 100 µmol L -1 final concentration). Slurries were sacrificed at approximately 0, 5 and 10 h after substrate addition by injection of 250 µL zinc chloride solution through the septum of exetainers.

Analytical methods
Analysis of 15 N composition of N2 (and any nitrous oxide: N2O) was determined by gas-chromatography isotope ratio mass spectrometry (GC-IRMS). A helium head space was introduced to filled exetainers and gas samples were manually injected as described in Dalsgaard et al. (2013). Any 15 N-N2O was reduced in a reduction oven and measured as 15 N-N2. Determination of 15 N in NH4 + was carried out by conversion of NH4 + to N2 with alkaline 260 hypobromite iodine solution (Risgaard-Petersen et al., 1995;Füssel et al., 2012). Ammonium was extracted from sediment in slurry and whole-core samples by shaking for 1 h with 2M KCl (1:1 sample:KCl) before any NH4 + analysis. The isotopic composition of the produced N2 was determined using a GC-IRMS as above. Recovery efficiency of 15 NH4 + following the hypobromite conversion was >95 %.

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For determination of total NH4 + , samples were extracted with KCl as above and NH4 + concentrations were analysed colorimetrically using the salicylate-hypochlorite method (Bower and Holm-Hansen, 1980).

Data calculations
Anammox and DNRA were detectable in slurry incubations, although both processes only played a minor role in NO3reduction at most sites. However, they may have interfered to a minor degree with the IPT calculations. Thus 270 areal rates of benthic N cycling processes were calculated according to Song et al. (2016) at all sites. The relative contribution of anammox to N2 production (ra) in slurries was calculated as in Song et al. (2013) using the average mole fraction of 15 NH4 + in the total NH4 + pool (FA) as this was demonstrated to increase linearly over time.
Fluxes of NO3and NH4 + were calculated using gradients (~0-1 cm and ~0-5 cm, respectively) of sediment pore water depth profiles and Fick's first law of diffusion. Porosity values were taken from the average porosities of 275 the integrated depth horizons and diffusion coefficients from Schulz (2006).

Sediment accumulation rates
Freeze-dried sediment samples for Strömmen were measured for 210 Pb by direct gamma counting using a high purity germanium detector (Ortec GEM-FX8530P4-RB) at Lund University. 210 Pb was measured by its emission at 46.5 keV. Self-absorption was measured directly and the detector efficiency was determined by counting a 280 National Institute of Standards and Technology sediment standard. Excess 210 Pb was calculated as the difference between the measured total 210 Pb and the estimate of the supported 210 Pb activity as given by 214 Pb ( 210 Pbexc = 210 Pbtotal − 214 Pb). Sediment accumulation rates for the four study sites were estimated by fitting a reactive transport model (Soetaert and Herman, 2008) to the 210 Pb depth profiles accounting for depth dependent changes in porosity (Sup. Fig. 2).

Pore water profiles
The O2 penetration depth is deepest (18 mm) at Ingaröfjärden, while at the other three sites the O2 penetration depth is relatively shallow (<4 mm; Table 2; Sup. Fig. 3). The diffusive uptake of O2 is high at Strömmen and Baggensfjärden (~14 mmol m -2 d -1 ) and low at Ingaröfjärden (3 mmol m -2 d -1 ; Table 2). All four sites are 290 characterized by a shallow sulfate methane transition zone (SMTZ), with near complete SO4 2removal between 7 and 15 cm (Fig. 4). Concentrations of CH4 increase with depth at all sites and are highest at Erstaviken (up to 8 mmol L -1 ) and lowest at Ingaröfjärden (max. ~2 mmol L -1 ). At Strömmen, Baggensfjärden and Erstaviken, H2S concentrations increase rapidly with depth below 2 cm, while at Ingaröfjärden this is observed below 10 cm. After a distinct maximum (of up to 1.3 mM in Ingaröfjärden), H2S concentrations decrease again with depth, and even 295 reach values close to zero at Strömmen and Erstaviken (at approximately 20 and 40 cm, respectively). The flux of H2S towards the sediment surface is high at all sites (~4 to 8 mmol m -2 d -1 ).
Dissolved Fe 2+ concentrations show a maximum directly below the sediment-water interface at all sites, with the highest maximum values at Strömmen (~60 µmol L -1 ), and a rapid decrease to values around zero in the upper centimeters of the sediment. At Strömmen and Erstaviken dissolved Fe 2+ concentrations increase again when H2S 300 is depleted at depth. At all sites, concentrations of HPO4 2and NH4 + are low near the sediment-water interface, and then increase with depth, first quickly then more gradually. Only at Strömmen HPO4 2decreases below ~15 cm. Bottom water NO3concentrations decrease from the inner archipelago towards the outer archipelago, i.e.
Strömmen > Baggensfjärden > Erstaviken > Ingaröfjärden. For the three most inshore sites NO3concentrations in the bottom water are higher than NO3concentrations in the sediments. In contrast, at Ingaröfjärden NO3concentrations in the surface sediments are almost four times higher than NO3concentrations in the bottom water.

Solid phase profiles
Sediment Corg concentrations are relatively high at all four sites (Fig. 5), whereas CaCO3 concentrations are low (< 3 wt. %; Table 4). Surface sediments are enriched in Corg by 1-2 wt. % when compared to sediments at depth.
Concentrations of Corg are highest at Strömmen and decrease from the inner archipelago towards the outer 310 archipelago (Table 4; Sup. Fig. 4). Sediment C/N ratios are somewhat lower in the top centimeters and become constant with depth. Overall C/N values decrease towards the outer archipelago. At all four sites, surface sediments are enriched in P. The thickness of this enriched surface layer ranges from 2 to 4 cm. At Strömmen, surface P concentrations are twice as high (ranging up to 165 µmol g -1 ) as those observed at the other sites. Below this enriched surface layer, P concentrations are mostly rather constant at all sites (ranging from 30 to 40 µmol g -315 1 ). Similar to the high concentrations in the surface layer at Strömmen, sedimentary P concentrations are also high at depth (40 to 50 µmol g -1 ), and two additional enrichments in P are observed at depth.
As a result of the relatively large enrichment in P in the surface sediments, Corg/Ptot is low in the surface sediment. At depth Corg/Ptot values are around the Redfield-ratio (Table 4). With the exception of Strömmen, surface sediments are enriched in Fe-oxides. This enrichment is most pronounced at Ingaröfjärden. At depth, Fe-oxide 320 concentrations are relatively constant and similar for all four sites. Just below the surface, between ~1 to 10 cm, a pronounced enrichment in FeS is observed. Only at Ingaröfjärden such a pronounced enrichment in FeS is not observed, and FeS is entirely absent above 2.5 cm. Pyrite concentrations are relatively low in the surface sediments and gradually increase with depth. At Ingaröfjärden, a peak in FeS2 is observed between 5 and 10 cm, superimposed on the gradual increase in FeS2.

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At all sites, Fe-bound P dominates the P in the surface sediments ( Fig. 6). At Strömmen, Fe-bound P remains an important fraction of solid phase P, also at depth, while for the other sites Fe-bound P only represents ~10-20 % of total P. Exchangeable P shows trends similar to those observed for Fe-bound P, but concentrations are low.
Detrital P, Authigenic P and P in organic matter all show relatively constant concentrations with depth. Only the P in organic matter is slightly enriched in the surface sediments. Below the Fe-bound P-dominated surface 330 sediments, P in organic matter is the largest fraction, representing between ~30 and 40 % of the total P and between ~30 and 50 % of reactive P (i.e., the sum of Fe-bound P, exchangeable P, P in organic matter and authigenic Ca-P). Authigenic Ca-P represents ~25 to 30 % and detrital P ~20 to 25 % of total P.

Benthic nitrogen cycling
Bottom water NO3concentrations decrease from Strömmen (17.8 µmol L -1 ) toward Ingaröfjärden (5.6 µmol L -1 , 335 Table 5). The flux of NH4 + out of the sediment also decreases seawards. The sediment acts as a weak source of NO3to the overlying water at Strömmen while it is a NO3sink at the other three sites (Table 5).
Denitrification is the major NO3 --reducing process at all sites ( Fig. 7; Table 5). Denitrification rates (Fig. 7) are highest at Strömmen (~1700 µmol m -2 d -1 ) and decrease towards the outer archipelago with the lowest rates at Ingaröfjärden (~100 µmol m -2 d -1 ). Nitrous oxide is not an important end-product of denitrification in whole core 340 incubations. Nitrification is the main source of NO3to at all sites, accounting for 60-89 % of all NO3supply (Table 5; Fig. 8). The importance of nitrification as NO3source relative to water column NO3increased towards the outer archipelago. DNRA was measurable but is not a significant NO3 --reducing pathway at any of the sites investigated, accounting for less than 1.5 % of total NO3reduced. Anammox plays only a minor role in overall N removal (< 1 % N2 produced) at the three inner archipelago sites but accounts for 33% of N2 production at 345 Ingaröfjärden (44.1 µmol m -2 d -1 ) where overall N2 production is lowest and heterotrophic denitrification was most limited in organic C substrate. Rates of N removal by denitrification are positively correlated with bottom water NO3concentrations and with organic carbon content (Fig. 9).  (Gidhagen, 1987), which at many locations in the Stockholm Archipelago leads to hypoxia (Karlsson et al., 2010;Conley et al., 2011). In addition to nutrient availability, spring bloom intensity and water depth, hydrological restriction may contribute to low O2

Phosphorus dynamics in a eutrophic coastal system
conditions. This is also reflected at our study sites, with Baggensfjärden being the most O2 depleted and restricted 360 basin (i.e. land-locked with narrow and relatively shallow connections to adjacent basins) and Ingaröfjärden being the least restricted and subsequently, the most consistently well-oxygenated basin throughout the year (Table 1; Figs. 1, 2 and Sup. Fig. 1).
High dissolved O2 concentrations allow the formation and presence of Fe-oxides (Fig. 5) in the surface sediments that bind P (e.g. Slomp et al., 1996;Fig. 6). Low dissolved O2 concentrations, however, lead to the dissolution of Fe-oxides in the surface sediments. The P associated with these Fe-oxides can then be released into the water column again. This mechanism leads to P recycling in basins with strong (seasonal) contrasts in bottom water redox conditions, such as Baggensfjärden, where the sediments are a sink for P in the winter and a source for P in the spring and the summer (Fig. 3c), as also described previously for other basins in the Stockholm Archipelago (Walve et al., 2018). Nevertheless, in year-round well-oxygenated basins, such as Ingaröfjärden, this seasonal P 370 recycling is (nearly) absent (Fig. 3a). In such basins, deeper O2 penetration, which might partly be related to the presence of macrofauna (Sup. Fig. 3), leads to a thicker Fe-oxide bearing layer (Fig. 5) and a larger and stable Febound P pool (Fig. 6), hence a larger enrichment of P in the surface sediments (Fig. 10). Besides Fe-oxides, a major part of the surface sediment P pool consists of P in organic matter (Fig. 6), which, based on the C/N values close to the Redfield-ratio (Fig. 5), is predominantly of marine origin. Part of the organic matter (and the P 375 associated with it) is lost with depth (Fig. 6), because the most labile organic matter is degraded in the upper centimeters of the sediment, releasing the P associated with it to the pore water. For our study sites in the Stockholm Archipelago we calculated that this surface sediment P pool, i.e. the P active in turn-over as earlier already suggested by Rydin et al. (2011), varies between 0.036 mol P m -2 at Baggensfjärden and 0.172 mol P m -2 at Ingaröfjärden (between ~1 and 5 g P m -2 , respectively; Fig. 10; Table 6). This is comparable to values found 380 for previously studied sites in the Stockholm Archipelago (1 to 7 g P m -2 ; Rydin et al., 2011;Rydin and Kumblad, 2019). The surface sediment P pool, could, however, have been much larger for Strömmen, Baggensfjärden and Erstaviken if all of the FeS in the surface sediments would seasonally transform to Fe-oxides. The lack of such a transformation is likely linked to the high upward flux of H2S to the surface sediment (4.2 to 7.6 mmol m -2 d -1 ; Table 3). Besides the H2S flux, there is a relatively large efflux of NH4 + from the sediments into the bottom water 385 (up to 1.4 mmol m -2 d -1 ; Table 5). Both the H2S and the NH4 + flux originate from decomposing organic rich sediments at depth (Fig. 4). Upon aerobic oxidation, two moles of O2 are consumed per mole of H2S or NH4 + (e.g. Reed et al., 2011). Thus, the oxygen demand resulting from these H2S and NH4 + fluxes is very high when compared to the diffusive flux of O2 into the sediment (3.1 -13.8 mmol m -2 d -1 ; Table 2). As a consequence of the high H2S flux, FeS is formed and/or preserved (Fig. 5), and formation of a large(r) pool of Fe-oxides and Fe-bound P pool 390 is hindered.

Phosphorus burial
Absolute P concentrations in the sediments in the Stockholm Archipelago (Figs. 6 and 10 in this study and in Rydin et al., 2011) are high (~30 to 50 μmol g -1 ) in comparison with most other studied sites in the coastal zone of the Baltic Sea (generally <30 μmol g -1 ; Jensen et al., 1995;Carman et al., 1996;Lenstra et al., 2018). The 395 relatively low Corg/Ptot values in the top ~2 cm, which are around the Redfield-ratio (Fig. 5), show that the seasonal O2 depletion of bottom waters in our study area is not severe or long enough to cause substantial preferential regeneration of P relative to C (Algeo and Ingall, 2007;Sulu-Gambari et al., 2018). The combination of high absolute P concentrations and relatively high sedimentation rates leads to relatively high rates of P burial (Table   6; Sup. The constant concentrations of most P forms in the sediment below the clearly "enriched" surface sediments, suggest there generally is little to no sink-switching of sediment P forms in the Stockholm Archipelago. The 410 curved shape of the porewater HPO4 2profiles indicate, however, that there is still some release of P to the porewater at depth and we attribute this to slow degradation of organic matter. Both the detrital and authigenic (Ca-P) fractions are likely buried in the form in which they reached the sediment-water interface. The general dominance of P in organic matter and apatite (authigenic and detrital P; Fig. 6) at depth (representing permanent P burial, since the release of P from organic P is only minor), agrees with previous findings for organic rich 415 sediments in the Baltic Sea (e.g. Jensen et al., 1995;Carman et al., 1996;Mort et al., 2010;Rydin et al., 2011).
By contrast, in the Bothnian Sea, Fe-bound P is a much more important P pool at depth (e.g. Egger et al., 2015; Lenstra et al., 2018). Evidence for potential sink-switching is only found at Strömmen, which is characterized by a larger Fe-bound P pool at depth (Fig. 6). This larger Fe-bound P pool at depth contributes to the high P burial rate at Strömmen (Table 6; Sup. Fig. 5). Coastal sediments with a shallow SMTZ, relatively high inputs of Fe-12 oxides and organic matter and high sediment accumulation rates are prime locations for formation of vivianitetype minerals (e.g. Slomp et al., 2013;Egger et al., 2015). The presence of dissolved Fe 2+ and decreasing dissolved HPO4 2concentrations at depth at Strömmen (Fig. 4) in combination with elevated Fe-bound P in the lower part of the record (Fig. 6), hence may result from the formation of a vivianite-type mineral.

Spatial differences in benthic N dynamics
Denitrification is by far the dominant pathway of NO3reduction at our study sites, accounting for ~80 to 99 % of total dissimilatory NO3reduction (DNRA + anammox + (2 x denitrification)) at the time of sampling (Table 4).
The dominant role of denitrification in removing N and the gradient from inner to outer archipelago agrees well with regional models based on long-term monitoring data, which show the highest N-removal capacity in the inner the inner archipelago, where Strömmen is located, annually removes approximately 3-5 times more N (~8-12 t N km -2 yr -1 ) than the intermediate and outer archipelago sites (~1-3 t N km -2 yr -1 ). Denitrification rates of both Baggensfjärden and Erstaviken are within this range (~2.5 and ~3 times lower than at Strömmen, respectively).
However, despite Ingaröfjärden being located in a basin adjacent to Erstaviken (Fig. 1) and modelled as having 435 an almost identical area-specific N retention capacity (Almroth-Rosell et al., 2016), denitrification rates were almost 20 times lower than those at Strömmen and ~8 to 6 times lower than at Baggensfjärden and Erstaviken, respectively. As such, N removal rates between adjacent basins may be more variable than assumed by models.
The differences in rates are likely related to lower organic matter inputs and subsequent lower sediment respiration rates as indicated by deeper O2 penetration at Ingaröfjärden (Table 2; Sup. Fig. 3). Suspended particulate organic 440 matter may also be removed more quickly from Ingaröfjärden due to its more direct connection to the open Baltic Sea (Fig. 1) permitting more rapid water exchange and transport of particulate organic matter out of the basin than at Baggensfjärden and Erstaviken (Engqvist and Andrejev, 2003).

Controls on benthic NO3reduction
Given the minor contributions of anammox and DNRA in these sediments at the time of sampling, we focus 445 predominantly on the control of heterotrophic denitrification in Stockholm Archipelago sediments at the time of sampling. Heterotrophic denitrification in sediments is limited by both the availability of NO3and Corg.
In sediments, NO3is supplied from overlying water and/or from nitrification in the surface layers (coupled nitrification-denitrification ;Seitzinger, 1988;Seitzinger et al., 2006). The relative importance of the two NO3sources to denitrification in coastal systems can be highly variable between locations and seasons (e.g. Seitzinger Jäntti et al., 2011;Bonaglia et al., 2014). We observed a distinct positive correlation between rates of denitrification and bottom water NO3concentration (Fig. 9) indicating a high capacity of the sediments to reduce riverine NO3loads along the seaward gradient, as shown for other coastal systems of the Baltic Sea (Asmala et al., 2017). We additionally demonstrate that benthic nitrification provided the major proportion (~55-90 %) of NO3which was reduced in the sediments at all four sites (Table 5; Fig. 8), as has been demonstrated in previous 455 studies and syntheses on coastal systems (e.g. see Seitzinger et al., 2006) and studies within the Baltic Sea (e.g. Silvennoinen et al., 2007;Bonaglia et al., 2014;Bonaglia et al., 2017;Hellemann et al., 2017). One of the highest contributions of nitrification to NO3production for denitrification (~85 %) was measured at Ingaröfjärden. At this site, the lowest overall denitrification rates and bottom water NO3concentrations were measured, despite the deep (18 mm) oxygen penetration providing a large sediment volume for nitrification to occur (Table 2). This high 460 oxygen penetration may in part be due to less Corg inputs and thus a lower Corg content (Table 2), discussed in section 4.2.1 and further below.

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Inputs of Corg provide both a C-source for heterotrophic processes (e.g. denitrification) as well as a source of NH4 + (from remineralisation processes) for nitrification and subsequent NO3production. In coastal sediments Corg is not thought to limit denitrification. However, in complex basin systems such as the Stockholm Archipelago, and 465 the Baltic Sea coastal zone in general, differences in ventilation and retention times between basins may mean that Corg inputs are more variable than assumed (see section 4.2.1). Available Corg in Ingaröfjärden (Table 2) may be less labile than at other sites due to such hydrological variations, with the deep (18 mm) oxygen penetration indicating a lower organic matter reactivity and sediment respiration compared to the other sites. Lower labile Corg availability will limit heterotrophic denitrification and may explain why anammox, an autotrophic process, is 470 more dominant at this site (Table 5; Fig. 7). The presence of the invasive polychaete Marenzelleria (Table 2) may also reduce N removal at Ingaröfjärden and enhance the efflux and transport of NH4 + from sediments (e.g. Hietanen et al., 2007;Bonaglia et al., 2013)

Seasonal cycles of N processes
Sampling and experiments in this study were carried out in late winter (March), a period in the Baltic Sea when the water column is well mixed, with cold and well oxygenated bottom waters and with persistently low organic inputs to sediments. However, conditions are of course not static throughout the annual cycle. Seasonal warming, stratification, phytoplankton blooms and consumption and release of nutrients as seen in year-round monitoring 480 data ( Fig. 3d; Sup. Fig. 1) will have marked effects on sediment nutrient cycling. Year-round bottom water monitoring data collected at Bäggensfjärden show that NO3accumulates annually in bottom waters during the autumn and winter months before being consumed during spring and summer by phytoplankton blooms (Fig. 3d).
Hypoxic bottom waters develop over summer following bloom collapse and subsequent enhanced deposition of fresh organic matter and enhanced benthic respiration during summer and early autumn. Bottom water total N 485 concentrations increase during summer in connection with the hypoxic events (Fig. 3d) due to enhanced benthic remineralization and subsequent NH4 + efflux from sediments.
Increased organic inputs following the spring bloom are likely to lead to increases in denitrification as the season progresses, as is commonly observed in coastal sediments (e.g. Piña-Ochoa and Álvarez-Cobelas, 2006;Jäntti et al., 2011;Bonaglia et al., 2014). Thus, a similar scenario would be assumed for the Stockholm Archipelago as for 490 other estuaries, leading to higher rates of denitrification during spring and early summer and a reduction in autumn and winter as organic inputs subside (e.g. Bonaglia et al., 2014). Depending on the bloom intensity and organic matter inputs during spring, increased benthic respiration may lead to more reduced conditions in surface sediments as bottom water O2 is depleted. The availability of NO3also declines under hypoxic/anoxic conditions due to NO3consumption in the water column, lower oxygen penetration and thus a reduced volume of surface sediment where nitrification can occur and from the reduced efficiency of nitrification under low oxygen conditions. The resulting high C/N conditions may cause process dominance to shift from N removal by denitrification (or anammox) to retention by DNRA (e.g. An and Gardner, 2002;Burgin and Hamilton, 2007;Giblin et al., 2013;Algar and Vallino, 2014;Kraft et al., 2014), as has been repeatedly demonstrated in field, laboratory and model studies (An and Gardner, 2002;Algar and Vallino, 2014;Kraft et al., 2014;van den Berg 500 et al., 2016;Kessler et al., 2018). Thus, under hypoxic conditions in summer/autumn, DNRA may become the dominant NO3 --reducing process, altering the role of sediments from a NO3sink through N2 production, to a source via increased NH4 + release by DNRA.
While we have not assessed NO3 --reducing process over different seasons at these four sites, we have demonstrated the microbial metabolic potential for DNRA is present through the detection of DNRA activity in 505 incubations at all four sites (Table 5). We suggest that it is highly likely that DNRA contributes to NH4 + efflux at sites during sporadic bottom water hypoxia. Thus, the capacity for N removal by denitrification may be reduced during bottom water hypoxia while the likelihood of N recycling by DNRA increases as shown in previous Baltic Sea studies (e.g. Jäntti et al., 2011;Jäntti and Hietanen, 2012;Bonaglia et al., 2014).

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Continued decreases in nutrient inputs to the Baltic Sea Andersen et al., 2017) and the Stockholm Archipelago (Karlsson et al., 2010) are likely to reduce phytoplankton growth, lead to reduced organic matter input into the sediments and, eventually, to higher O2 concentrations in bottom waters.
Our results indicate that increases in bottom water O2 would likely impede the observed present-day P recycling pattern at the seasonally hypoxic sites (Fig. 3c), allowing thicker Fe-oxide bearing layers and a larger Fe-bound P 515 pool in the surface sediments (e.g. Slomp et al., 1996), hence a larger (semi-permanent) surface sedimentary P pool. This process will, however, be delayed due to the prior deposition of organic rich sediments which results in a high upward flux of H2S (Table 3), i.e. legacy of hypoxia hindering the formation of Fe-oxides that can bind P. Because of this legacy effect, we expect that artificial reoxygenation of bottom waters (e.g. Stigebrandt and Gustafsson, 2007), if applied in the Stockholm Archipelago, is unlikely to be a long-term effective measure 520 towards improving the water quality since it does not stimulate permanent P burial in these sediments and a large impact on the Fe-P pool is hindered by the high upward H2S flux. Further nutrient reduction for the Stockholm Archipelago is expected to eventually lead to a reversal from export of P to the open Baltic Sea to import of P from the open Baltic Sea (Savchuk, 2005;Almroth-Rosell et al., 2016). This shows that improvement of the water quality in the Stockholm Archipelago is to a great extent coupled to nutrient management strategies for the entire 525 Baltic Sea.
Our results indicate that in the Stockholm Archipelago, N likely goes through cycles of retention and removal throughout the year in relation to bottom water hypoxia. N is removed by denitrification during colder months when NO3availability is high, while DNRA is likely to increase during hypoxic, NO3 --depleted months.
Reductions in the frequency of hypoxic bottom waters will thus reduce the amount of time that sediments 530 potentially recycle bioavailable N via DNRA and sediments may be more likely to act as a net sink for N through denitrification on an annual basis.
Continued recovery of the Stockholm Archipelago is also likely to lead to (re-)colonisation by bioturbating macrofaunal populations that have been driven out by hypoxic bottom waters (Diaz and Rosenberg, 2008;Voss et al., 2011). This may enhance P burial and denitrification by sediment reworking and oxygenation (e.g. Pelegri

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and Blackburn, 1995;Laverock et al., 2011;Norkko et al., 2012). While we still lack the predictive capabilities required to allow us to assess how fauna may influence sediment biogeochemistry (Griffiths et al., 2017; recolonization by fauna at inner archipelago sites is likely to maintain and reinforce active P and N removal processes. Thus, these coastal sediments are likely to continue to contribute to removal of P and N as long as we 540 continue to actively reduce nutrient inputs.

Conclusion
Seasonally hypoxic sites in the Stockholm Archipelago are characterized by active sedimentary P recycling, because low bottom water O2 concentrations seasonally destabilize Fe-oxides that bind P in the surface sediments.
A high upward flux of H2S, due to prior deposition of organic rich sediments in a low O2 setting, leads to the 545 formation and preservation of FeSx instead of burial of Fe-oxides at these sites. At the site where bottom waters are well-oxygenated year round, the surface sedimentary P pool is mainly characterized by P bound to Fe-oxides and organic matter, in a pool that is 5 times larger than that at the most hypoxic site (~0.172 versus ~0.036 mol P m -2 ). At depth, sedimentary P is dominated by P in organic matter and apatite. Only for the site in the inner Archipelago (Strömmen), there is an indication for sink-switching, i.e. authigenic formation of a vivianite-type 550 Fe(II)-P mineral, at depth. Burial rates of P at our sites in the Stockholm Archipelago are high (0.03-0.3 mol m -2 y -1 ) because of the combined effect of high sediment accumulation rates and high sedimentary concentrations of P. Benthic denitrification is the primary NO3 --reducing pathway in the Stockholm Archipelago leading to remediation of NO3introduced from the water column and from benthic nitrification. Decreases in denitrification rates follow the gradient of bottom water NO3and sedimentary Corg content from the inner archipelago towards 555 the open Baltic Sea from ~1700 to ~100 µmol N m -2 d -1 . Combining our process measurements with available monitoring data, it is likely that N in the Stockholm Archipelago undergoes seasonal cycles of removal through denitrification/anammox and recycling by DNRA. Further reductions in P and N inputs are expected to reduce the frequency of hypoxic events and to continue to support the Stockholm Archipelago's capacity to remove P and N loads.

Code and data availability
Monitoring data are available from the Swedish Meteorological and Hydrological Institute (SMHI, 2019). All other data, if not directly available from the tables and supplement, will be made available in the PANGAEA database. In the meantime data is available upon request to the authors.

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The supplement related to this article is available online at:

Author contribution
NvH, ER, DC, and CS designed the research. NvH, ER, MH, CH, WL and CS carried out the fieldwork. NvH, ER, MH, JK and WL performed the analyses. All authors interpreted the data. NvH, ER and CS wrote the paper with comments provided by DC, MH, CH, JK and WL.    Fig. 3) Table 3. Diffusive fluxes of pore water H2S.

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denitrification' indicates the proportion of denitrification supported by NO3from nitrification (as opposed to water column NO3 -). Bottom water NO3concentrations and NH4 + and NO3fluxes from the surface sediments into the water column (calculated from pore water profiles), including standard error (SE).