An interestuarine comparison for ecology in TIDE

Following processes (Sanders et al. 1997, Statham et al. 2011) occur in both the water column and sediments:

• export to adjacent coastal oceans;
• biological uptake, might be in resistant forms within the sediment;
• losses to the atmosphere by nitrification and denitrification processes;
• organic matter mineralisation.

Although most is transported to the adjacent coastal zone, in some cases removal by production of gaseous forms can be significant. Benthic denitrification has been shown to remove 44 % of nitrate in estuaries. Biological uptake depends on primary production and has been show to have significant effect on dissolved inorganic nitrogen concentrations in spring (Arndt et al. 2011). Nitrification and denitrifcation are influenced by high organic matter loads, nitrate concentrations and the presence of an oxic or sub-oxic zone (Statham et al. 2011). In the Elbe the turbidity maximum zone is observed together with a peak in nitrification. The ammonium source for nitrification rather came from particulate organic nitrogen degradation, than it came from dissolved organic nitrogen degradation in the turbidity maximum zone. Schlarbaum et al. showed that adsorption and desorption processes could play an important role in the distribution of dissolved organic nitrogen and ammonium concentrations (Schlarbaum et al. 2010 and refs herein). Although, nitrogen is thought to be not very particulate reactive opposed to phosphorus (Statham et al. 2011). The rate of denitrification reaches a maximum at concentrations of 500 µM nitrate (7 mg/l) (Jickels et al. 2000). Denitrification and consequent nitrogen removal is generally promoted by low oxygen concentrations (Soetaert et al. 2006). Although, nitrogen is less likely to be removed in deeper channels (Alexander et al. 2000). The larger the sediment surface, the larger the potential oxic-anoxic boundary zone for denitrification (Dähnke et al. 2008). Both intertidal marshes and mudflats also have been demonstrated to be important in nitrogen retention (Mortimer et al. 1998, Van Damme et al. 2009). Higher influx of nitrate was observed on intertidal mudflats in the freshwater part of the estuary (Mortimer et al. 1998). Fluxes by the sediment-water exchange can be enhanced by macro-faunal activity through bio-irrigation (Mortimer et al. 1998, Braeckman et al. 2010).

Inputs for inorganic nitrogen species are associated with atmospheric exchange (N-deposition), surface water run-off, rivers and ground water input. Input by rivers is of primary importance. Either source can be direct, or indirect (through organic matter input). The importance of groundwater input is largely unknown.

Dissolved organic compounds are not considered, but are possibly of more importance than assumed at present (Statham et al. 2011).

Phosphorus

Phosphorus is regulated by inorganic and biological interactions. Hence, phosphorus dynamics are dependent on the presence of different particulate and dissolved phosphorus species and biological activity. Phosphorus is required for photosynthesis, general metabolism, cell wall synthesis and energy transfer (e.g. ATP) (Statham et al. 2011). Phosphate and particulate inorganic phosphorus bound to suspended matter both seem to be the most abundant phosphorus species involved in phosphorus dynamics (van der Zee et al. 2007).

Phosphate is very particle reactive. In estuaries a buffering mechanism exists at the suspended particulate matter-water and sediment-water interface.

In turbid estuaries (SPM > 50 mg/l), phosphate is strongly adsorbed to iron-oxyhydroxides (fig 3). As particulate phosphorus arrives in the high salinity zone, competition for adsorption sites by stronger anions (OH-, F-, SO42-, …), causes suspended matter to release phosphate to the water column again (van Beusekom & Brockmann 1998, Deborde et al. 2007, Statham et al. 2011).

Organic matter mineralisation for dissolved inorganic phosphate is a rather slow process, and its importance is largely dependent on residence times (69 days for the Gironde). Nevertheless, for the Gironde example less than 10 % is mineralized in the sediment while about 50 % is mineralised in the estuarine turbidity maximum zone (Deborde et al. 2007). Most organic matter mineralisation appears to occur within the turbidity maximum zone, usually to be found in the lower salinity reaches (Abril et al. 1999). The higher mineralisation rate can be attributed to the larger reactive surface area of the suspended particulate matter and the availability of more degradable organic matter opposed to higher salinity reaches (van Beusekom & Brockmann 1998).

Furthermore the residence time (and thus time for biogeochemical transformation processes) of non-conservative behaving suspended matter is significantly longer than for conservative behaving dissolved substances (Statham et al. 2011).

The suspended particulate matter might also settle before it arrives in the high salinity zone and form a pool of phosphate buried in the sediment. The phosphate burial in the sediment is very dependent on the redox state. When the iron-oxyhydroxides in the sediments are reduced, phosphate is again released to the water column. Also, when sediments are exported more downstream, iron-sulfate complexes are formed preferential upon iron-phosphate complexes, and phosphate is released (Statham et al. 2011). Combined with re-suspension phosphate release can be further enhanced (van der Zee et al. 2007). However, some phosphate may be more permanently buried when precipitated as authigenic minerals like apatites (van Beusekom & Brockmann 1998).

In general, phosphorus retention is promoted by high oxygen concentrations, opposite to the requirements for nitrogen retention (Soetaert et al. 2006). Tidal mudflats studied in the Humber, showed very small and variable phosphate fluxes comparable to fluxes in other temperate estuaries (Mortimer et al. 1998 and refs herein). Fluxes by the sediment-water exchange can be enhanced by macro-faunal activity through bio-irrigation (Mortimer et al. 1998, Braeckman et al. 2010).

Rivers are the most principal pathways for phosphorus input. Atmospheric inputs can be considered negligible compared to atmospheric deposition for nitrogen. Groundwater inputs are largely unknown and can be function of bottom geo-morphology (e.g. if a so-called ‘iron curtain’ is present, groundwater input can be inhibited). Phosphate dynamics are difficult to generalize and are very dependent on the physico-chemical properties of the environment (Statham et al. 2011).

Phosphate and particulate inorganic phosphorus (adsorbed to SPM) often are the main phosphorus species involved in phosphate dynamics (van der Zee et al. 2007). However, also dissolved organic phosphorous appears to become of more interest in phosphorus dynamics, but remains largely unstudied up to now (Statham et al. 2011).

Silica

Dissolved silica is mainly regulated by biology along the estuarine gradient. Thus fluxes are highly seasonal dependent. It is a critical component for diatoms to build up their cell walls (frustules). Hence, in summer lowest dissolved silica concentrations and corresponding highest biogenic silica (opal) concentrations are observed. The residence time has to be sufficiently long for the diatoms to take up the dissolved silica. Highest concentrations are observed in winter and behave conservatively along the mixing gradient. Diatoms have been shown to take up 50 to 70 % of the total silica load during the productive period (Carbonnel et al. 2009). When algal growth is inhibited, or stopped (dissolved silica limitation < 0.01 mM ~ 0.3 mg DSi/l, light, grazing), diatom frustules might sink and be buried in the sediment. Biogenic silica dissolution increases with salinity and some bacterial communities have been shown to increase dissolution too. Furthermore, phytoliths within macrophytes (e.g. Reed) can be added by topsoil erosion or directly from vegetation of riverbanks or tidal marshes to the biogenic silica pool. Tidal marshes are most likely very important silica recyclers to the estuary. Retention of biogenic silica in tidal marshes, originating from imported biogenic silica of diatoms via suspended matter sedimentation and phytoliths in Reed, combined with increased residence times and increased sediment-water interactions, intensifies biogenic silica dissolution processes. Tidal marshes have been demonstrated to be a source for dissolved silica when freshwater discharge and dissolved silica concentrations within the estuary are low (Struyf et al. 2006, 2007). Recycling capacity has been shown to be higher for young marshes than for older, more elevated marshes. This can be related to different sedimentation rates (Struyf et al. 2007). In intertidal mudflats silica fluxes were found very small within the Humber estuary. However, this might not be as representative, since primary production within this estuary is very low. However, low dissolution patterns could also be linked to rather low overlying dissolved silica concentrations (Mortimer et al. 1998). Fluxes by the sediment-water exchange are in general enhanced by macro-faunal activity through bio-irrigation (Mortimer et al. 1998, Braeckman et al. 2010). However, most dissolved silica originates from rock weathering and can be associated with riverine input (Carbonnel et al. 2009, Statham et al. 2011). In rivers dissolved silica concentrations usually remained high all year round and biogenic silica low (Carbonnel et al. 2009), except sometimes in larger rivers (e.g. Elbe, personal communication Andreas Schöl 2013). In the tidal river dissolved silica concentrations are almost completely consumed, while biogenic silica concentrations rise. Contribution of phytoliths in the Scheldt to detrimental biogenic silica was found non-significant. Hence, delivery of silica from Reed phytoliths could not account for the increase in biogenic silica during the productive period. In summer dissolved silica concentrations can be consumed up to limitation levels (Van Damme et al. 2005, Soetaert et al. 2006). Within the estuary biogenic silica is mainly delivered by diatoms, especially in summer. In winter, most is imported as dissolved silica (Carbonnel et al. 2009). When silica export is inhibited, nuisance algal blooms can appear downstream in the coastal zone (Rousseaux et al. 2006). Opposed to the summer bloom, the spring bloom had no significant effect upon silica dynamics (Carbonnel et al. 2009). To understand regulating mechanisms, it is important to also consider biogenic silica fluxes (Carbonnel et al. 2009). Dynamics in the freshwater zone drive the dynamics in the coastal zone (Rousseaux et al. 2006).

Nutrient ratios

The ideal nutrient ratio for diatom growth is defined by Redfield as 106/16/1/16 corresponding to C/N/P/Si respectively (Billen & Garnier 2007). When light climate and residence time are suitable, primary production will increase with nutrient concentrations rising. However, within most estuaries light is limiting (SPM > 10 mg/l, mixing depth) and change in nutrient ratios will only affect algal growth there where a better light climate is reached, e.g. in the estuarine plume of the Humber estuary, beyond Spurn point (Nedwell et al. 2002, Soetaert et al. 2006). Last decades, nutrient loads have increased, because of population increase and land use change. Combined with unequal water treatment efficiency, this has lead to disproportional nutrient ratios in many estuaries (Statham et al. 2011). Nutrient ratios define phytoplankton species composition and succession along the estuarine gradient; hence define ecological functioning (Carbonnel et al. 2009). Phosphorus is typically limiting in freshwater, while nitrogen is limiting in estuarine and coastal systems. However, large riverine nitrogen input and more efficient phosphorus removal have shifted the limitation in many northern temperate estuaries to a limitation for phosphorus (Sanders et al. 1997, Soetaert et al. 2006). Nitrogen is mostly related to diffuse input sources and therefore much more difficult to manage. Although to a lesser extent, also silica loads have been altered by human activities. Hydrological measures such as embankments have reduced the contact with the riparian vegetation and consequently lowered silica input in the estuary. Furthermore, deforestation has lowered the baseflow delivery of silica twice to threefold in temperate European watersheds (Struyf et al. 2010). Hence, lowered silica input and increasing nitrogen and phosphorus loads have increased overall silica limitation in estuaries.