Quantity and quality of organic matter (detritus) drives N2 effluxes (net denitrification) across seasons, benthic habitats, and estuaries

Authors


Abstract

[1] N2 flux rates (net denitrification) were measured over a diel cycle, seasonally, in 12 benthic habitats across three warm temperate Australian coastal systems. Dark N2-N fluxes were strongly controlled by sediment oxygen demand (SOD) across the 3 estuaries, 4 seasons, and 12 benthic habitats (r2 = 0.743; p < 0.001; n = 142; slope = 0.0170). However, some of the slopes differed significantly between seasons and among estuaries and habitats, and all of the slopes were correlated with the δ13C values and C:N ratios of sediment organic matter. Ternary mixing diagrams with the contribution of algal, seagrass, and terrestrial/mangrove material to sediment organic matter showed that habitats, seasons, and estuaries dominated by a mixture of seagrass and algal material had the lowest slopes, and slopes increase as habitats, seasons, and estuaries have an increasing contribution from terrestrial/ mangrove material. Overall, the slopes of dark N2 fluxes versus SOD were low compared to previous studies, most likely due to either, or a combination of, the C:N ratio of the organic matter, the mixture of C:N ratios making up the organic matter, the structure of the organic matter, and/or the SOD rates. This study demonstrated that it is not only the quantity but also the type (quality), and maybe the mixture, of organic matter that is an important control on denitrification. As such, rapid global changes to detrital sources to coastal systems due to losses of mangrove, seagrasses, and saltmarshes, and associated increases in algae and macrophytes, are also expected to impact system level losses of nitrogen via denitrification.

1 Introduction

[2] Global change is rapidly altering the type, and function, of primary producers in coastal ecosystems. Mangroves, seagrasses, and saltmarshes have declined by around 30 to 50% over the past few decades due to reclamation, deforestation, and urbanization (reviewed in Mcleod et al. [2011]). Over the same period, there has been a large increase in phytoplankton and ephemeral macrophyte production (eutrophication) in many coastal systems due to nutrient over-enrichment [Cloern, 2001; McGlathery et al., 2007], and an associated increase in the ratio of pelagic to benthic production [Borum and Sand-Jensen, 1996; Ferguson and Eyre, 2010]. The phenology of phytoplankton blooms, and associated quality of phytodetritus, has been altered due to climate changes [e.g., Nixon et al., 2009]. Combined, these changes in primary producers will significantly effect the quantity, and quality, of nonliving (detrital) organic matter in coastal systems, which in turn impacts ecosystem structure and function [Kelaher et al., 2013]. However, it is not fully understood how these changes in detrital resources influence key ecosystem processes such as denitrification.

[3] Denitrification is a critical ecosystem process that permanently removes nitrogen from an ecosystem by converting fixed nitrogen to di-nitrogen gas, which can then be lost to the atmosphere. Nitrogen lost to the atmosphere as di-nitrogen gas acts as a control on system level primary productivity. The considered importance of denitrification is demonstrated by the numerous denitrification studies in most ecosystems [Seitzinger et al., 2006]. These studies have identified several primary factors that control denitrification, including the supply of labile organic matter and nitrate, bottom water oxygen concentrations, and several secondary factors such as the presence or absence of macrofauna, macrophytes, benthic microalgae, H2S, and FeS [Cornwell et al., 1999; Canfield et al., 2005]. In sediments where the overlying water is well oxygenated and has low nitrate concentrations, a supply of labile carbon is probably the most important controlling factor on denitrification. As such, the carbon and nitrogen cycles of costal ecosystems are closely linked via denitrification, which requires a source of organic matter (detritus) to proceed, but can limit the production of organic matter via nitrogen removal.

[4] Although many studies have looked at the effect of changes in organic matter on benthic denitrification in coastal systems, these have mostly focused on organic matter quantity, not quality, in response to understanding the effects of the deposition of excess organic matter production (phytodetritus) to the sediments due to nutrient over-enrichment. For example, several experimental studies have shown that denitrification is enhanced when low C:N organic matter (e.g., phytoplankton, glucose, or yeast) is added to sediments [Brettar and Rheinheimer, 1992; Caffrey et al., 1993; Fulweiler et al., 2008; Fulweiler et al., 2013], although coupled nitrification-denitrification, and associated denitrification efficiency, can be reduced [Caffrey et al., 1993; Eyre and Ferguson, 2009], and Banks et al. [2013] found no response in benthic denitrification to added low C:N organic matter in sediments already enriched in organic matter. Several studies have also identified a positively linear relationship between dark rates of benthic denitrification and sediment oxygen demand (SOD; a proxy for the quantity of organic matter oxidation; Table 1). Seitzinger and Giblin [1996] found a slope of 0.116 for a compilation of benthic denitrification rates and SOD for continental shelf sediments where the source organic matter was low C:N phytodetritus. Similarly, flowthrough reactor experiments using permeable carbonate sands and seawater, with phytoplankton as the organic matter source, had a SOD versus dark N2 efflux slope of 0.114 [Santos et al., 2012]. Organic matter oxidation would increase the supply of NH4+ from ammonification for coupled nitrification-denitrification, increase the availability of electron donors for denitrification, and modify the sediment redox conditions.

Table 1. Slope of the Relationship Between Sediment Oxygen Demand (µmol O2 m−2 h−1) and Dark N2 Efflux (µmol N2-N m−2 h−1; Denitrification)a
Sediment TypeSlopeSource
  1. aAll regressions were forced through zero as there would be no N2 efflux when respiration equals zero.
  2. bData from Figure 4 in Glud et al. [2008] and Figure 3 in Eyre et al. [2008].
  3. cData from Figure 5 in Eyre et al. [2011a]; r2 is slightly less than in Eyre et al. [2011a] due to forcing through zero.
  4. dSlopes calculated from Figure 1 in Oakes et al. [2011].
Subtropical to temperate continental shelf muds0.116Seitzinger and Giblin [1996]
Subtropical to temperate muds and sands0.086Fennel et al. [2009]
Very coarse permeable carbonate sands0.033bEyre et al. [2008]
Temperate muds, range of habitats from subtidal shoals to seagrass0.036 to 0.107 (0.063 average)Piehler and Smyth [2011]
Subtropical muds, range of habitats from subtidal shoals to seagrass0.036cEyre et al. [2011a]
Fine muddy sand, low C:N organic matter0.129dOakes et al. [2011]
Fine muddy sand, high C:N organic matter0.022dOakes et al. [2011]
Very coarse permeable carbonate sands, low C:N organic matter0.114Santos et al. [2012]
Fine to very coarse permeable carbonate sands, high C:N organic matter0.019 to 0.047 (0.035 average)Eyre et al. [2013]
Fine to very coarse permeable carbonate sands, low C:N organic matter0.089Eyre et al. [2013]

[5] Much less is known about the effect of organic matter quality on benthic denitrification in coastal systems. The only experimental study to compare additions of organic matter of different quality on benthic denitrification rates found that high C:N organic matter added to fine muddy sands in a temperate climate suppressed denitrification, most likely due to competition for nitrogen by heterotrophs processing the refractory organic matter [Oakes et al., 2011]. Similarly, Fulweiler et al. [2013] speculated that the decrease in benthic denitrification during a 214 day long mesocosm experiment was due to the organic matter becoming more refractory (high C:N) resulting in an increased competition for nitrogen. These conditions would have been an advantage to sulphate reducers that fix their own nitrogen and may also suppress nitrification, both of which would decrease N2 effluxes [Fulweiler et al., 2013]. Nitrate uptake in the dark (inferred to be denitrification) following additions of low and high C:N organic matter to marine sediments found that the magnitude of the uptake was not related to the C:N ratio but was related to the lability (structure) of the organic matter [Dahllof and Karle, 2005].

[6] A number of previous studies in coastal systems have looked at SOD versus dark N2 efflux slopes. The effect of high C:N organic matter on denitrification rates can be seen in SOD versus dark N2 efflux slopes in the experiment of Oakes et al. [2011], which were 0.129 for low C:N (7.2) organic matter (phytodetritus) and 0.022 for high C:N (28.2) organic matter. A compilation of 657 measurements of denitrification and SOD from a range of aquatic systems found a slope of 0.086 (Table 1) and suggested that the lower slope than that identified by Seitzinger and Giblin [1996] was simply due to the larger data set [Fennel et al., 2009], although it may have been due to the mixture of organic matter (i.e., high and low C:N organic matter) across the different aquatic systems. In another compilation study across multiple estuaries, seasons, and habitats, the average slope of dark rates of denitrification versus SOD was 0.063 (Table 1), and it was suggested that the lower slope compared to the earlier work of Seitzinger and Giblin [1996] was due to weaker coupling between nitrification and denitrification [Piehler and Smyth, 2011], but again it may have been due to the mixture of organic matter (i.e., high and low C:N organic matter) across the different habitats. This same study also found a range of different slopes in different seasons and habitats ranging from 0.036 to 0.107 (Table 1). Similarly, in a subtropical coastal system, a low slope of 0.036 was found across multiple benthic habitats [Eyre et al., 2011a]. A recent compilation of dark denitrification rates and SOD measurements in permeable carbonate sands found two slopes, and it was argued that the difference was due to the type of organic matter being mineralized [Eyre et al., 2013]. A steeper slope of 0.089 was proposed to be driven by the mineralisation of episodic inputs of low C:N phytodetritus, whereas a lower slope of 0.036 was driven by inputs of high C:N organic matter from coral reefs (Table 1).

[7] Previous studies that have looked at the control of SOD on denitrification have all used rates measured in the dark. Benthic denitrification rates measured in shallow coastal systems in the light are usually different to dark rates due to the influence of benthic primary producers [Eyre and Ferguson, 2005; Ferguson and Eyre, 2013]. Benthic production during the light can increase the penetration of oxygen into the sediments with contrasting effects. Coupled nitrification-denitrification may be enhanced if NH4+ is readily available [Risgaard-Petersen et al., 1994; An and Joye, 2001; Eyre and McKee, 2002], whereas denitrification driven by water column NO3 may decrease due to consumption of NO3 by benthic microalgae and an increase in the diffusional path length [Risgaard-Petersen et al., 1994]. It is unknown if light denitrification rates will be well correlated with SOD due to the influence of benthic primary producers.

[8] The first hypothesis of this study is that dark rates of benthic N2 efflux (net denitrification) will be strongly positively correlated with SOD, but the slope of this relationship will vary depending on the sources (quality) of organic matter driving respiration (SOD). We tested this hypothesis by undertaking benthic N2 efflux and SOD measurements (data in Maher and Eyre [2011]) over four seasons, in up to 12 benthic habitats, in three estuaries on the east coast of Australia. This gave a broad range of organic matter (detritus) quality. Stable isotopes and sediment molar C:N ratios were used to identify changes in the sources of organic matter driving benthic respiration. The second hypothesis of this study is that SOD will not be a good predictor of light rates of benthic denitrification due to the overriding control of benthic production. We tested this hypothesis by making the above benthic N2 efflux measurements over diel cycles. The findings from this study give insight into how changes in detrital resources in coastal systems associated with global change will influence key ecosystem processes such as denitrification.

2 Methods

2.1 Study Area

[9] Detailed descriptions of the three study estuaries can be found in Eyre and Maher [2010] and Maher and Eyre [2011]. Briefly, the estuaries are located along the southeast Australian coast and fall along an estuarine maturity gradient where they have evolved by gradually infilling [Roy et al., 2001]. Wallis Lake is an immature stage estuary with a large central mud basin, Camden Haven is an intermediate stage estuary with a more restricted shallower central basin, and the Hastings River Estuary is at a mature stage with a river-dominated system characterized by river channels with a highly restricted/absent central mud basin. These different stages result in distinct differences in estuarine area, water residence time, catchment area, and freshwater inflow [see Maher and Eyre, 2011, Table 1], the distribution of benthic habitats in each system (see maps in Eyre and Maher [2010]), and ecosystem-scale carbon cycling [Maher and Eyre, 2012]. A total of 12 benthic habitats was identified in the three estuaries using a combination of underwater video, diving transects, and remote sensing techniques (Table 2). The benthic habitats were delineated based on their depth (subtidal, intertidal), sediment grain size (mud dominant, sand dominant), geomorphology (channels, depositional basins, shoals), and dominant autotrophs (macroalgae, seagrass (Halophila ovalis, Zostera capricorni, Posidonia australis, and Ruppia megacarpa)).

Table 2. Benthic Habitats, Descriptions, and Abbreviations (Modified From Maher and Eyre [2010])
HabitatDescriptionAbbreviation
Intertidal sandIntertidal habitats composed primarily of sandIS
Intertidal mudIntertidal habitats composed primarily of mud/siltIM
Subtidal sandsSubtidal shoal habitats composed of sandSS
Subtidal mudsSubtidal shoal habitats composed of mud/siltSM
Fluvial muds and sandsChannel habitats in the middle/upper estuary dominated by fluvial sands and/or mudsFMS
Marine channelChannel habitats, lower estuary, clean quartz sandsMC
Deep subtidal mudsSubtidal habitats >4 m deep dominated by mudsDM
ZosteraSeagrass habitats with Zostera capricorni dominantZ
HalophilaSeagrass habitats with Halophila australis dominantH
RuppiaSeagrass habitats with Ruppia megacarpa dominantR
MacroalgaeHabitats dominated by macroalgaeMA
PosidoniaSeagrass habitats with Posidonia australis dominantP

[10] Rainfall was below average, water temperature varied from ~13°C to ~ 27°C, and bottom water oxygen concentrations generally ranged from ~5 mg L−1 to ~10 mg L−1 during the study period, although values as low as 1 mg L−1 were recorded at dawn in some seagrass areas [Maher and Eyre, 2011]. The estuaries have a relatively small catchment population (24,000–48,000) [Maher and Eyre, 2011] although there is modification of the catchment for agricultural uses (for full description of estuarine stressors, see Eyre and Maher [2010]).

2.2 Benthic N2 Flux (Net Denitrification) Measurements

[11] Flux sites that best represented the benthic habitat within each estuary were selected using benthic habitat maps [Eyre and Maher, 2010]. A total of 11, 16, and 17 sites was sampled in triplicate over four seasons in the Hastings River (n = 132), Camden Haven (n = 192), and Wallis Lake (n = 204) estuaries, respectively, during field campaigns in July 2006 (winter), October 2006 (spring), February 2007 (summer), and April 2007 (autumn).

[12] N2 fluxes were measured using a combination of in situ benthic chamber incubations and field laboratory-based sediment core incubations. Benthic chambers were used in habitats characterized by seagrass and/or large burrowing macrofauna, as the important structural elements of these habitats could not be captured in cores. Cores were used in all other habitats. We believe that there is no large systematic bias introduced by using a combination of cores and chambers because of the high r2 between benthic fluxes measured in cores and chambers and measures which are independent of the incubation device such as sediment δ13C (see section 3). In addition, previous studies have also shown that in shallow water coastal systems, cores and chambers give comparable nutrient flux data [Asmus, 1986].

[13] Details of the core and chamber incubations are given in Maher and Eyre [2011]. Briefly, triplicate sediment cores were collected at each site with plexiglass tubes (95 mm ID × 500 mm length) pushed approximately 200 mm into the sediment, leaving an overlying water volume of ~1.9 L. Cores were discarded if there was any surface disturbance of sediments. Uncapped sediment cores were placed in one of four large (150 L) perspex tanks filled with site water. Stirring was maintained at a speed just below the resuspension threshold via rotating magnets throughout the equilibration and incubation periods. Light was delivered by four high pressure 400 W sodium bulbs (Philips SON-T AGRO) with in situ light conditions emulated by shading individual cores to site-specific mean daily irradiance levels (±5%). Following a 24 h equilibration period, the cores were incubated over a 24 h diel cycle.

[14] Dissolved oxygen concentrations (Hach LDO optode DO meter, ±0.01 mg L−1) and pH and temperature (Denver AP25 pH probe, ±0.001 pH units, ± 0.1°C) were measured, and alkalinity and N2:Ar samples were collected at 0 (i.e., dusk) and 10 or 12 h during the dark cycle and 0 (i.e., dawn) and 12 or 14 h during the light cycle. Alkalinity samples were withdrawn with a plastic syringe and 0.45 µm filtered water was transferred to a 10 mL Milli-Q soaked and sample-rinsed polyethylene vial for alkalinity analysis. To minimize the introduction of bubbles, N2 samples were collected in duplicate by allowing water to flow, driven by the reservoir head, directly into 7 mL gas-tight glass-stoppered glass vials filled to overflowing. The replacement water for all samples was withdrawn from a sealed collapsible reservoir bag, also equilibrated at in situ light (±5.0%) and temperature (±1°C) conditions, to maintain constant Ar concentrations. Alkalinity samples were refrigerated at 4°C. N2:Ar samples were poisoned with 20 μL of 5% HgCl2 and stored submerged at ambient temperature. N2:Ar samples were analyzed on a membrane inlet mass spectrometer with oxygen removal [Eyre et al., 2002]. The O2, alkalinity, and TCO2 (alkalinity and pH) data are presented in Maher and Eyre [2011]. At the conclusion of the core incubations, the top 0–10 mm depth was sectioned for sediment TOC, TN, and δ13C concentrations. Samples were placed in aluminium foil and frozen (−20°C) until analysis. For chamber incubations, a sediment core was collected from within each chamber (95 mm ID, 200 mm deep) and was sampled as above. Full details of analytical methods are presented in Maher and Eyre [2010]. The sediment TOC and TN concentrations and δ13C values are summarized in Maher and Eyre [2010] and are presented here in full.

[15] Benthic chambers of a similar design to those described by Webb and Eyre [2004], but with a larger volume to surface area ratio (50 L:840 cm2), were used for in situ diel incubations in macrophyte (i.e., seagrass and macroalgae) communities and shallow subtidal/intertidal habitats colonized by burrowing macrofauna (predominantly the thalassinidean shrimp Trypaea australiensis). Chambers were equilibrated for ≥24 h with the lids open, allowing free exchange of water. Chambers were then sealed at dusk and sampled for O2, pH, temperature, N2:Ar, and alkalinity (as per core incubations) over an 18 h incubation (0 h, dawn, dawn + ~6 h) with the light period capturing solar noon.

2.3 Calculations

[16] N2 fluxes across the sediment-water interface were calculated using the start- and end-point concentration data, corrected for the addition of replacement water, as a function of incubation time, core/chamber water volume, and surface area. Dark flux rates were calculated using concentration data from the nighttime part of incubations and light flux rates were calculated using concentration data from the daytime part of incubations. Careful consideration was given to the possible effect of bubbles on N2 fluxes [Eyre et al., 2002]. All samples over 96% O2 saturated were excluded from N2 flux calculations because it was clear that these samples were affected by bubbles. As such, many dark and light rates are missing due to both the initial and end dusk samples being over 96% O2 saturated. Bubbles have the effect of reducing N2 concentrations, giving much higher dark N2 flux rates and lower light N2 flux rates or apparent rates of N fixation. Despite the missing N2 fluxes, both dark and light net denitrification rates were determined in all seasons for all estuaries, and in all habitats in all estuaries, but not all habitats in all seasons in all estuaries. Because N2 fluxes were measured, they include both canonical denitrification and anammox. N2 fluxes reflect the balance between N fixation and denitrification and, as such, are a measure of net denitrification. The terms N2 efflux and net denitrification are used interchangeably.

2.4 Ternary Mixing Diagrams

[17] Ternary mixing diagrams were used to determine the relative contribution of algae, seagrass, and terrestrial/ mangrove organic matter to the sediment organic matter in each season in each estuary as a whole, and in each habitat across all estuaries. However, instead of using a single end-member value, a range of average measured δ13C values from the three systems and single literature molar C:N (N:C) ratios were used for each end-member. N:C ratios were used in graphs for ease of visualization. The algal δ13C end-member ranged from −18.6 ± 3.8‰ (n = 73) for bulk live microphytobenthos and phytodetritus to −14.8 ± 1.2‰ for epiphytes [Maher and Eyre, 2010] with a C:N ratio of 6.6 (N:C = 0.15). The bulk live seagrass end-member ranged from −15.0 ± 2.9‰ (n = 18) for Zostera capricorni to −11.21 ± 0.32‰ (n = 6) for Posidonia australis [Maher and Eyre, 2010] with a C:N ratio of 20 (N:C = 0.05) [Gonneea et al., 2004]. The bulk live terrestrial/ mangrove end-member ranged from −32.4 ± 2.0‰ (n = 18) [Maher and Eyre, 2010] to −27.0 ± 4.0‰ for terrestrial organic matter [Michener and Schell, 1994; Marshall et al., 2007] with a C:N ratio of 50 (N:C = 0.02) [Gonneea et al., 2004].

2.5 Statistical Analysis

[18] A paired t test was used to determine if there was a statistical difference between dark and light fluxes of N2 (α = 0.05). One-way analyses of variance were used to explore seasonal, habitat type, and estuary differences for N2 fluxes (dark, light, and net) and sediment TOC, TN, and δ13C. Where significant differences were found (i.e., p < 0.05), Tukey post hoc tests were used to determine homogenous subsets. All analyses were done using SPSS v20.0 software.

[19] To test for significant differences between individual sediment oxygen demand versus dark N2 effluxes slopes, the t statistic was calculated as a function of the difference of the slopes divided by the difference in the standard error of the slopes, and p was calculated using the t value and n − 4 degrees of freedom [Kleinbaum et al. 2008].

3.0 Results

[20] N2 fluxes ranged from 3.5 µmol N2-N m−2 h−1 in the SS habitat in Wallis Lake in the dark during autumn to 551.6 µmol N2-N m−2 h−1 in the SM habitat in Hastings River in the dark during summer. N2 fluxes were significantly higher in the dark than the light (p = 0.003) (Table 3). Dark N2 fluxes were significantly different between seasons (p = 0.002), estuaries (p < 0.026), and habitats (p < 0.001) (Table 3). N2 fluxes were higher in spring and winter than in summer and autumn. N2 fluxes in the Hastings River Estuary and Wallis Lake were higher than in Camden Haven. The H, IS, Z, MA, and R habitats had higher rates of net denitrification than the SS, SM, DM, P, MC, IM, and FMS habitats. Light N2 fluxes were only significantly different between habitats with the IS and MA habitats having higher rates than the R, P, Z, and IM habitats which, in turn, had higher rates than the MC, H, and FMS habitats. The lowest net denitrification rates in the light were in the SM and SS habitats.

Table 3. Seasonal Dark and Light N2 Fluxes (µmol N2-N m−2 h−1) in the 12 Habitats Across Wallis Lake, Camden Haven, and the Hastings River Estuary (Mean ± SE, n = 2 or 3)a
HabitatWinterSpringSummerAutumn
DarkLightDarkLightDarkLightDarkLight
  1. aND = many dark and light rates are missing due to both the initial and end dusk samples being over 96% O2 saturated.
Hastings River
FMSNDND17.817.825.8 ± 8.647.3 ± 10.398.3 ± 25.565 ± 16.3
HNDNDNDND65.3ND19.1 ± 18.513.1 ± 5.5
IMND100.2 ± 13.748.7 ± 0.3ND89.8 ± 65.6ND35.6ND
IS89.9ND172.3 ± 118.5157.6 ± 45.6212.6 ± 9.3173.7 ± 29.3234.2257 ±
MCND74 ± 9.3NDND25.113.1 ± 842.359.8 ± 7.5
SM18.9 ± 12.9NDNDNDND5.5 ± 2.18.4 ± 2.9ND
SSNDNDNDND16.2 ± 3.128.4 ± 1113.5 ± <0.121 ± 11.4
ZNDND279.9ND445.2 ± 23.4ND247.5 ± 48.9ND
Camden Haven
FMSND13.25.3 ± 3.18.258.9 ± 17.953 ± 32.816.32.4
HNDNDNDNDND12.3ND21.4
IMNDNDND83.3 ± 15.3227.4ND75.0189.6
ISNDND11.1 ± 7.9ND40.3NDNDND
M ISNDND87.9 ± 20.717 ± 9.8NDNDNDND
M ZNDND118.8NDNDNDNDND
MCNDND26.221.2 ± 15.820.0 ± 5.245.6 ± 38.19.3 ± 4.113.8
RNDNDNDND206.129.2 ± 15.586.6 ± 1.8147.8 ± 38.4
SM32.9 ± 6.526.5 ± 7.531.4 ± 4.013.8 ± 4.244.9 ± 10.225.7 ± 10.24.97.8 ± 4.7
SSNDND30.74.7 ± 0.313.718.314.518
Z135.5 ± 24.2NDND118.1ND75.8102.7 ± 15.1ND
Wallis Lake
DMNDND17.28.6 ± 3.9ND11.6 ± 5.9ND15.6 ± 5.8
FMSND31.811.5 ± 5.28.1 ± 3.18.8 ± 0.137.3 ± 13.853.5 ± 4.6ND
HNDND62ND121.8 ± 7.275.1 ± 28.2119.2 ± 32.0ND
IMNDND11.9 ± 8.061.811.4ND21.7 ± 13.310.9
ISNDND19.9ND6.7 ± 3.7ND12.7 ± 6.2ND
MA34.7NDNDNDND205.6 ± 66.7199.4ND
MCNDND0.112.212.6 ± 1.943.5NDND
P78 ± 40.455.1ND93.8 ± 12.350.6ND188.333
RNDNDNDND244.7 ± 38.268.1 ± 38.9132.1 ± 24.8281.6 ± 57.9
SMNDND7.28.312.9 ± 1.14.313.6ND
SSNDNDNDNDND7.53.5ND
ZNDNDNDND150.9 ± 24.1ND203.6 ± 6178.4

[21] Sediment δ13C values ranged from −27.3‰ in the SS habitat during winter in the Hastings River Estuary to −11.4‰ in the SS habitat during summer in Wallis Lake (Table 4). Sediment δ13C values were significantly different between seasons (p = 0.02), estuaries (p < 0.001), and habitats (p < 0.001) (Table 4). In summer and spring, sediment δ13C values were more enriched than in winter and autumn. Sediment organic matter in Wallis Lake had more enriched δ13C values than in Camden Haven, which in turn, had more enriched δ13C values than in the Hastings River Estuary. The MA habitat had the most enriched sediment δ13C values, followed by the P and R habitats, then the Z and H habitats, and then the SM, DM, IM, IS, SS, and MC habitats. The FMS habitat had the most depleted sediment δ13C values.

Table 4. Seasonal Sediment δ13C, Total Organic Carbon, and Total Nitrogen Concentrations (%) in the 12 Habitats Across Wallis Lake, Camden Haven, and the Hastings River Estuarya
HabitatWinterSpringSummerAutumn
δ13CTOCTNδ13CTOCTNδ13CTOCTNδ13CTOCTN
  1. aHabitat abbreviations given in Table 2.
Hastings River
FMS−25.61 ± 0.510.83 ± 0.210.06 ± 0.01−24.14 ± 0.320.48 ± 0.120.05 ± 0.01−22.83 ± 0.770.26 ± 0.090.05 ± 0.01−23.38 ± 0.340.44 ± 0.090.05 ± 0.01
HNDNDNDNDND0.1ND1.95 ± 0.140.16 ± 0.01ND2.01 ± 0.180.13 ± 0.01
IM−24.23 ± 0.61.07 ± 0.150.07 ± 0.01−23.24 ± 0.492.28 ± 1.350.15 ± 0.06−21.52 ± 1.240.79 ± 0.110.08 ± 0.01−22.46 ± 0.10.75 ± 0.120.06 ± 0.01
IS−24.42 ± 1.020.29 ± 0.090.02 ± 0.01−18.76 ± 0.320.22 ± 0.040.03 ± 0.01−19.08 ± 0.180.45 ± 0.150.04 ± 0.01−18.99 ± 0.320.46 ± 0.190.04 ± 0.02
MC−22.91 ± 0.573.67 ± 0.040.24−23.87 ± 0.114.88 ± 0.650.25−19.98 ± 0.43.36 ± 0.020.26−23.95 ± 0.043.06 ± 0.060.2
SM−26.88 ± 0.830.4 ± 0.130.02 ± 0.01−24.15 ± 0.240.72 ± 0.120.05 ± 0.01−21.92 ± 0.310.58 ± 0.070.05 ± 0.01−22.72 ± 0.11.86 ± 0.250.12 ± 0.01
SS−27.28 ± 0.70.21 ± 0.040.02 ± 0.01−20.91 ± 0.570.2 ± 0.030.02−24.07 ± 0.850.2 ± 0.030.02−21.84 ± 0.290.23 ± 0.040.03 ± 0.01
Z−20.57 ± 0.561.01 ± 0.190.07 ± 0.01−19.72 ± 0.330.77 ± 0.240.07 ± 0.02−19.74 ± 0.070.85 ± 0.050.08 ± 0.01NDNDND
Camden Haven
FMS−23.4 ± 0.287.51 ± 3.220.46 ± 0.14−23.81 ± 0.652.9 ± 0.520.16 ± 0.04−21.67 ± 0.492.16 ± 0.750.12 ± 0.03−23.21 ± 0.484.62 ± 0.770.23 ± 0.03
HND4.07 ± 0.10.38 ± 0.01ND3.92 ± 0.190.35 ± 0.02ND3.01 ± 0.110.27 ± 0.01 4.28 ± 0.080.32 ± 0.03
IM−15.43 ± 0.063.08 ± 0.170.3 ± 0.02−15.79 ± 0.161.61 ± 0.40.17 ± 0.04ND2.02 ± 0.740.2 ± 0.07−15.81 ± 0.112.35 ± 0.230.23 ± 0.02
IS−21.62 ± 0.180.42 ± 0.060.04−21.29 ± 0.240.240.03ND0.5 ± 0.090.05 ± 0.01−17.65 ± 0.560.69 ± 0.040.08 ± 0.01
M IS−22.26 ± 0.740.42 ± 0.050.03−20.27 ± 0.720.82 ± 0.190.06 ± 0.01ND0.64 ± 0.020.06−21.42 ± 0.112.17 ± 0.350.15 ± 0.02
M Z−21.02 ± 0.321.05 ± 0.150.07 ± 0.01−20.03 ± 0.182.76 ± 1.080.2 ± 0.08ND2.42 ± 0.40.18 ± 0.03−20.46 ± 0.271.82 ± 0.140.14 ± 0.01
MC−21.93 ± 0.670.11 ± 0.010.01−21.44 ± 0.620.15 ± 0.030.02−21.210.12 ± 0.010.02−17.58 ± 0.280.130.02
R−15.81 ± 0.072.81 ± 0.520.28 ± 0.05−15.84 ± 0.562.42 ± 0.910.24 ± 0.08ND3.35 ± 0.180.34 ± 0.01−16 ± 0.041.64 ± 0.430.15 ± 0.04
SM−20.88 ± 0.722.83 ± 0.320.21 ± 0.03−20.77 ± 0.462.07 ± 0.230.15 ± 0.02−21.011.94 ± 0.210.19 ± 0.01−20.31 ± 0.632.49 ± 0.130.2 ± 0.01
SS−24.83 ± 0.920.1 ± 0.010.02−21.14 ± 0.870.13 ± 0.010.02−20.76 ± 0.080.17 ± 0.03NDNDNDND
Z−16.33 ± 0.054.04 ± 0.050.37−15.87 ± 0.174.44 ± 0.050.42ND4.04 ± 0.040.37 ± 0.01−16.4 ± 0.343.78 ± 0.470.28 ± 0.03
Wallis Lake
DM−16.72 ± 1.093.64 ± 0.350.31 ± 0.03−16.44 ± 0.693.4 ± 0.350.33 ± 0.04−16.62 ± 0.663.78 ± 0.350.36 ± 0.04−16.55 ± 0.633.34 ± 0.340.31 ± 0.03
FMS−21.45 ± 0.323.49 ± 0.170.24 ± 0.02−22.77 ± 0.113.82 ± 0.270.26 ± 0.02−22.56 ± 0.693.1 ± 0.310.27 ± 0.03−22.37 ± 0.183.64 ± 0.230.28 ± 0.02
HND3.76 ± 0.30.35 ± 0.03−13.982.96 ± 0.80.29 ± 0.08−13.66 ± 0.041.42 ± 0.40.14 ± 0.04NDNDND
IM−20.13 ± 1.321.12 ± 0.270.09 ± 0.02−21.61 ± 0.021.98 ± 0.130.17 ± 0.01−21.25 ± 0.12.3 ± 0.340.2 ± 0.03−21.32 ± 0.142.05 ± 0.150.17 ± 0.01
IS−19.79 ± 1.180.49 ± 0.170.05 ± 0.01−20.05 ± 0.580.7 ± 0.020.07−18.03 ± 0.590.36 ± 0.090.04 ± 0.01−18.71 ± 0.130.89 ± 0.120.08 ± 0.01
MA−13.85 ± 0.053.55 ± 0.180.33 ± 0.01−13.7 ± 0.072.97 ± 0.60.29 ± 0.04−13.85 ± 0.15.06 ± 0.230.47 ± 0.02NDNDND
MC−23.59 ± 1.36NDNDNDNDND−19.13 ± 0.50.07 ± 0.01ND−17.45 ± 0.930.07 ± 0.01ND
P−14.83 ± 0.71.03 ± 0.080.09 ± 0.01−13.86 ± 0.171.13 ± 0.210.11 ± 0.02−14.73 ± 0.591 ± 0.660.1 ± 0.06NDNDND
R−13.95 ± 0.053.73 ± 0.030.35−14.31 ± 0.254.08 ± 0.380.39 ± 0.04−14.19 ± 0.085.53 ± 0.140.52 ± 0.01NDNDND
SMNDND0.12 ± 0.01−22.25 ± 0.152.28 ± 0.110.17 ± 0.01−22.17 ± 0.112.92 ± 0.240.22 ± 0.02−21.95 ± 0.142.65 ± 0.110.19 ± 0.01
SS−15.97 ± 0.530.3 ± 0.050.03−13.37 ± 0.180.41 ± 0.020.05 ± 0.01ND0.41 ± 0.110.04−11.38 ± 0.170.33 ± 0.040.03
Z−16.94 ± 0.381.02 ± 0.210.08 ± 0.02−14.73 ± 0.055.59 ± 0.380.54 ± 0.04−14.35 ± 0.145.23 ± 0.160.49 ± 0.02NDNDND

[22] Sediment total organic carbon (TOC) concentrations ranged from 0.07% in the MC habitat during summer and autumn in Wallis Lake to 7.51% in the FMS habitat during winter in the Hasting River Estuary (Table 4). TOC concentrations were significantly different between estuaries (p < 0.001) and habitats (p < 0.001), but not seasons (p = 0.979). Sediments in Wallis Lake had the highest TOC concentrations, followed by Camden Haven, with the lowest sediment TOC concentrations in the Hasting River Estuary. The highest TOC concentrations were in the R and MA habitats, followed by the SM, DM, Z, H, and FMS habitats and then the P, MC, and IM habitats. The SS and IS habitats had the lowest TOC concentrations.

[23] The patterns of sediment nitrogen concentrations were similar to those for TOC (Table 4). Sediment nitrogen concentrations ranged from 0.02% in a number of habitats during different seasons in all systems to 0.54% in the Z habitat during spring in Wallis Lake. Sediment nitrogen concentrations were significantly different between seasons (p < 0.001), estuaries (p < 0.001), and habitats (p < 0.001). Sediments in Wallis Lake had the highest nitrogen concentrations, followed by Camden Haven, with the lowest sediment nitrogen concentrations in the Hasting River Estuary. The highest sediment nitrogen concentrations were in the R and MA habitats, followed by the SM, DM, Z, H, and FMS habitats and then the P, MC, and IM habitats. The SS and IS habitats had the lowest sediment nitrogen concentrations.

4.0 Discussion

4.1 Comparison of Denitrification Rates

[24] Few studies have looked at denitrification rates across multiple habitats within a coastal system [e.g., Piehler and Smyth, 2011], and only one of these has measured denitrification rates over a diel cycle [Eyre et al., 2011a]. Subtidal, intertidal, and seagrass habitats are the only common habitats across all these multiple habitat studies. Similar to this study, most multiple habitat studies have found higher dark rates of denitrification in seagrass habitats compared to intertidal and subtidal shoals, although the higher rates may only occur in some seasons [Eyre et al., 2011a; Eyre et al., 2011b; Piehler and Smyth, 2011; Smyth et al., 2012]. These more recent high rates of net denitrification using the N2:Ar technique (28 to 824 µmol m−2 h−1) contrast with earlier very low rates (<1 to 35 µmol m−2 h−1) of denitrification measured in seagrass communities using isotope pairing [e.g., Risgaard-Petersen et al., 1998; Risgaard-Petersen and Ottosen, 2000; Welsh et al. 2000]. Eyre et al. [2011a] suggested that such differences may reflect a difference between tropical and temperate systems, with higher rates in tropical systems and lower rates in temperate systems. However, Piehler and Smyth [2011] and Smyth et al. [2012] recently measured high rates of net denitrification in temperate seagrass communities using N2:Ar ratios. This suggests that the differences are either methodological, with lower rates measured using isotope pairing and higher rates measured using N2:Ar, or due to the biogeochemistry of the different seagrass species. For example, low rates of coupled nitrification-denitrification might reflect irregular oxygen release by the roots of some temperate seagrass species [Frederiksen and Glud, 2006]. Further work using both methods simultaneously in different seagrass communities is required.

4.2 Organic Matter Control on N2 Effluxes

[25] Dark N2 fluxes were strongly controlled by SOD (data from Maher and Eyre [2011]) across the 3 estuaries, 4 seasons, and 12 benthic habitats (r2 = 0.743; p < 0.001; n = 142; slope = 0.0170; Figure 1 and Table 5). Dark N2 fluxes also showed a strong correlation with dark TCO2 effluxes (r2 = 0.499; p < 0.0001; n = 142; slope = 0.0147; TCO2 data from Maher and Eyre [2011]) demonstrating that much of the SOD is due to organic matter oxidation. Average water column nitrate concentrations across most habitats were <3 µmol L−1 [Eyre and Maher, 2010], suggesting that much of the N2 flux was driven by coupled nitrification-denitrification. As such, the control of SOD on N2 effluxes most likely reflects an increased supply of NH4+ from ammonification for coupled nitrification-denitrification, the availability of electron donors for denitrification, and modification of sediment redox conditions.

Figure 1.

Dark N2 flux (µmol N2-N m−2 h−1) versus sediment oxygen demand (µmol O2 m−2 h−1) across the 3 estuaries, 4 seasons, and 12 benthic habitats.

Table 5. Slope of the Relationship Between Sediment Oxygen Demand (µmol O2 m−2 h−1) and Dark N2 Efflux (µmol N2-N m−2 h−1)a
 SeasonSloper2np
  1. aRegressions were forced through zero, as there would be no N2 efflux when respiration equals zero. ns = not significant.
Estuary
AllAll0.01700.743142<0.001
HastingsWinter0.02980.7124ns
 Spring0.02410.7456<0.05
 Summer0.02120.96417<0.001
 Autumn0.03120.75422<0.001
 All0.02370.82849<0.001
Camden HavenWinter0.02200.7807<0.05
 Spring0.02310.75317<0.001
 Summer0.02850.89616<0.001
 Autumn0.01480.90210<0.001
 All0.02140.73950<0.001
Wallis LakeWinter0.01300.6973ns
 Spring0.01140.71511<0.01
 Summer0.01320.96119<0.001
 Autumn0.01930.70219<0.001
 All0.01460.82052<0.001
Habitat
Intertidal sands (IS)All0.02410.69714<0.01
Intertidal muds (IM)All0.02410.80112<0.001
Subtidal sands (SS)All0.02540.7909<0.001
Subtidal muds (SM)All0.02540.44027<0.05
Fluvial muds and sands (FMS)All0.03320.53634<0.05
Marine channel (MC)All0.02440.7417<0.05
Deep subtidal muds (DM)All0.01880.9343ns
Zostera (Z)All0.02110.66112<0.01
Halophila (H)All0.01260.7179<0.01
Ruppia (R)All0.01420.8899<0.05
Macroalgae (MA)All0.01390.7123ns
Posidonia (P)All0.01650.7113ns

[26] There were differences however, in the slope of dark N2 fluxes versus SOD (benthic respiration) between seasons, between estuaries, and between habitats (Table 5). Although only some of the slopes were significantly different from each other (Tables 6 and 7), all the different slopes were highly correlated with δ13C values of sediment organic carbon across the different habitats (Figure 2a), and different seasons, in different estuaries (Figure 2b). The δ13C value of organic matter reflects its source [Peterson, 1999] and, as such, the δ13C values of sediment organic carbon reflect the mixture of different sources of organic matter within the sediments (see later discussion of changes during mineralization). This suggests that the slope of dark N2 fluxes versus benthic respiration is driven by the type, or mixture, of organic matter undergoing mineralization [Eyre et al., 2013]. An increase in the gradient of the slope was related to more depleted sediment δ13C values. Sediment C:N ratios were also correlated with the slope of dark N2 fluxes versus benthic respiration across both the different habitats (Figure 2c) and different seasons in different estuaries (Figure 2d). Interestingly, higher C:N ratios were related to an increase in the gradient of the slope. However, the C:N relationship was weaker than with sediment δ13C values, demonstrating that the type, or mixture, of organic matter undergoing mineralization was a more important control than its C:N ratio.

Table 6. The p Values for Test of Significance Between Sediment Oxygen Demand Versus Dark N2 Efflux Slopes for Individual Habitatsa
 DMFMSHIMISMAMCPRSMSSZ
  1. aValues in bold = <0.05. Abbreviations for habitats are given in Table 2.
DM            
FMS0.000           
H0.0040.000          
IM0.0810.0430.000         
IS0.1240.0570.0020.998        
MA0.2200.0000.6590.0170.024       
MC0.0530.0400.0000.9190.9260.019      
P0.5500.0010.2660.0840.1030.5820.077     
R0.0060.0000.2660.0020.0050.9170.0000.481    
SM0.0180.0730.0000.7180.7420.0030.7710.0310.000   
SS0.0050.0460.0000.6640.7030.0040.7150.0290.0000.999  
Z0.2460.0030.0000.3270.3890.0320.2200.2050.0000.1450.054 
Table 7. The p Values for Test of Significance Between Sediment Oxygen Demand Versus Dark N2 Efflux Slopes for Different Seasons in the Three Estuariesa
 Hastings WinterHastings SpringHastings SummerHastings AutumnCamden Haven WinterCamden Haven SpringCamden Haven SummerCamden Haven AutumnWallis Lake WinterWallis Lake SpringWallis Lake SummerWallis Lake Autumn
  1. aValues in bold = <0.05. H = Hastings, CH = Camden Haven, WL = Wallis Lake.
H Winter            
H Spring0.502           
H Summer0.1140.478          
H Autumn0.8490.1530.001         
CH Winter0.3500.6910.8000.036        
CH Spring0.3610.8180.4010.0210.782       
CH Summer0.3220.3210.0010.4090.0850.050      
CH Autumn0.0550.0420.0000.0000.0510.0010.000     
WL Winter0.1060.0690.0110.0000.0960.0100.0000.546    
WL Spring0.0230.0100.0000.0000.0090.0000.0000.0640.610   
WL Summer0.0250.0120.0000.0000.0110.0000.0000.1380.9510.239  
WL Autumn0.1520.2740.3190.0010.4490.1650.0010.0350.0690.0020.002 
Figure 2.

Sediment δ13C versus the slope of dark N2 flux (µmol N2-N m−2 h−1) versus sediment oxygen demand (µmol m−2 h−1) (a) across the 12 benthic habitats and (b) for each season in each estuary; molar C:N ratio versus the slope of dark N2 flux (µmol N2-N m−2 h−1) versus sediment oxygen demand (µmol O2 m−2 h−1) (c) across the 12 benthic habitats and (d) for each season in each estuary; and average sediment oxygen demand versus the slope of dark N2 flux (µmol N2-N m−2 h−1) versus sediment oxygen demand (µmol O2 m−2 h−1) (e) across the 12 benthic habitats and (f) for each season in each estuary.

[27] Ternary mixing diagrams showing the contributions of algal, seagrass, and terrestrial/ mangrove material to sediment organic matter give some insight into how changes in the type, or mixture, of organic matter influence the slope of dark N2 fluxes versus benthic respiration across different habitats, estuaries, and seasons (Figures 3 and 4). Sediments dominated by approximately equal mixtures of seagrass and algal material have the lowest slopes. The slopes increase for habitats that also have a contribution from terrestrial/mangrove material, with the steepest slope in the FMS habitat reflecting approximately equal proportions of terrestrial/mangrove and algal material. Similarly, the lowest slopes were in Wallis Lake where organic matter in the sediments was dominated by equal mixtures of algae and seagrass, and the highest slopes were in the Hastings River Estuary where terrestrial/mangrove material made a larger contribution (Figure 4). The slopes also increased in the seasons that had the largest contribution of terrestrial/ mangrove material, such as winter in the Hastings River Estuary, summer in Camden Haven and autumn in Wallis Lake (Figure 4). These seasonal interactions probably reflect differences in the delivery and trapping of terrestrial organic material due to different estuarine geomorphology.

Figure 3.

Ternary mixing diagram showing the relative contribution of algae, seagrass, and terrestrial/mangrove organic matter to each habitat. Also shown for each habitat is the slope of dark N2-N flux versus sediment oxygen demand (from Table 5). Habitat abbreviations given in Table 2. N:C ratios were used for ease of visualization. C:N ratio of 6.6 = N:C ratio of 0.15. C:N ratio of 20 = N:C ratio of 0.05. C:N ratio of 50 = N:C ratio of 0.02.

Figure 4.

Ternary mixing diagram showing the relative contribution of algae, seagrass, and terrestrial/mangrove organic matter in the three estuaries for each season. Also shown for each estuary for each season is the slope of dark N2-N flux versus sediment oxygen demand (from Table 5). C:N ratio of 6.6 = N:C ratio of 0.15. C:N ratio of 20 = N:C ratio of 0.05. C:N ratio of 50 = N:C ratio of 0.02.

4.3 Processes Driving Organic Matter Control on N2 Effluxes

[28] Although it is quite clear that the type, or mixture, of organic matter is driving the dark N2 fluxes versus SOD slopes across habitats, seasons, and estuaries (Figures 2a and 2b), the mechanisms involved were not entirely clear. It may be due to one, or a combination of, the C:N ratio of the organic matter, the mixture of C:N ratios making up the organic matter, the structure of the organic matter, and/or the overall SOD rates. All this mechanisms would influence the community composition of total bacteria, ammonia-oxidizing archaea, and denitrifiers, which have also been correlated to sediment δ13C in an estuarine system [Abell et al., 2013]. We will explore each of these mechanisms below.

[29] Overall, the dark N2 fluxes versus SOD slopes for this study were low (0.0114 to 0.0332; Table 5) compared to most previous compilation studies (Table 1). For example, a slope of 0.116 has been reported for continental shelf sediments [Seitzinger and Giblin, 1996], a slope of 0.086 was found for a compilation of 657 denitrification and SOD measurements [Fennel et al., 2009], and an estuarine study across multiple seasons and habitats found an average slope of 0.063 [Piehler and Smyth, 2011] (Table 1). Multiple lines of evidence suggest that higher slopes [e.g., Seitzinger and Giblin, 1996] are driven by the decomposition of low C:N phytodetritus and lower slopes (this study) may be driven by the decomposition of mostly high C:N sources of organic matter such as seagrass, mangrove, and terrestrial material. The lower slopes in Fennel et al. [2009] and Piehler and Smyth [2011] are probably also due to a mixture of organic matter sources driving denitrification in these studies, whereas phytodetritus would be the dominant source of organic matter to the continental shelf sediments that had the higher slope of 0.116 [Seitzinger and Giblin, 1996]. Flowthrough reactor experiments using permeable carbonate sands and seawater, with phytoplankton as the organic matter source, had a benthic respiration versus dark N2 efflux slope of 0.114 [Santos et al., 2012]. The only experimental study to compare additions of organic matter of different quality on dark N2 fluxes found a slope of 0.129 for low C:N (7.2) organic matter (phytodetritus) and a slope of 0.022 for high C:N (28.2) organic matter (slopes calculated from Figure 1 in Oakes et al. [2011]). Eyre et al. [2013] also speculated that high C:N organic matter from coral reefs may be driving low dark N2 fluxes versus benthic respiration slopes (0.036) in permeable carbonate sediment with higher slopes (0.089) driven by episodic input of low C:N phytodetritus.

[30] The dark N2 fluxes versus SOD slopes may also be driven by high C:N organic matter which would release less nitrogen as N2 for a given amount of respiration. A lack of N release would be enhanced due to competition for nitrogen by heterotrophic bacteria [Oakes et al., 2011]; N limitation of the microbial decomposition of high C:N organic material results in the uptake and accumulation of nitrogen by heterotrophic bacteria [Tupas and Koike, 1991; van Duyl et al., 1993; Lomstein et al., 1998]. Several types of bacteria can assimilate NH4+, including sulphate reducers and fermentative bacteria [Koike and Sumi 1989]. Additionally, sulphate reducers can fix nitrogen [Nielsen et al., 2001]. N2 effluxes are a measure of denitrification minus N fixation and as such, increased N fixation will result in a reduced N2 efflux. Furthermore, coupled nitrification-denitrification may be suppressed by H2S produced during sulphate reduction [Joye and Hollibaugh, 1995; Fulweiler et al., 2013]. Consistent with higher rates of heterotrophic N fixation during decomposition of higher C:N organic material is the reduction in dark N-fixation rates in permeable carbonate sands when low C:N phytodetritus was deposited [Eyre et al., 2008]. Nitrogen assimilation and N fixation by heterotrophic bacteria would be a nitrogen conservation process in high C:N seagrass communities that would otherwise lose large amounts of nitrogen via denitrification due to high rates of benthic respiration, i.e., the loss of nitrogen via denitrification per unit of respiration is an order of magnitude lower (Table 5) than the loss in a low C:N environment such as the continental shelf [Seitzinger and Giblin, 1996] (Table 1).

[31] If the C:N ratios were an important control on the SOD versus N2 efflux relationships in this study, it would be expected that higher slopes would be driven by low C:N algal material. However, the steeper slopes occur in sediments with higher amounts of high C:N terrestrial/mangrove organic matter, which is the opposite of what was expected. Because of the stronger relationship between the slopes and the δ13C than compared with C:N ratios, one explanation may be that the structure of the decomposing organic matter is influencing the slope (Figures 2a–2d). For example, Dahllof and Karle [2005] found that the C:N ratio of organic matter added to sediments was not an important control on nitrate uptake in the dark (inferred to be due to denitrification). Instead, they argued that nitrate uptake was related to the structure of the organic carbon, with greater nitrate uptake driven by organic matter containing easily degraded lipids and starch, and lower nitrate uptake driven by organic matter containing relatively refractory cellulose and lignin. This would explain the low dark N2 fluxes versus SOD slopes in the seagrass habitats and Wallis Lake, which is dominated by seagrass production [Maher and Eyre, 2011], but not the high slopes in the habitats and systems with higher inputs of terrestrial/mangrove organic matter.

[32] Alternatively, the dark N2 fluxes versus SOD slopes in this study may reflect different components of the high C:N organic matter or a mixture of low and high C:N organic matter. High C:N organic matter dominates within the study systems, giving the low slopes compared to previous studies, but the proportion of the labile component of the high C:N organic matter, or the low C:N proportion of the organic matter mixture, driving the dark N2 fluxes versus benthic respiration slopes may vary across the habitats, seasons, and estuaries. For example, the slopes increase for habitats with increasing amounts of terrestrial/mangrove material, but the SS and IS habitats that have higher proportions of algal material for a given δ13C values also have slightly higher slopes (Figure 3). Even more subtle is the P habitat, which has a slightly higher algal contribution and associated higher slope than the R and H seagrass communities for a given δ13C value (Figure 3). The sediment organic matter δ13C value and molar C:N ratio are a measure of the remaining organic matter after the more labile organic matter has been mineralized. This is consistent with more depleted sediment δ13C values associated with higher slopes, as removal of more labile 13C-enriched cellulose material results in more 13C depleted lignin-rich material remaining in the sediments [Benner et al., 1987].

[33] The type, or mixture, of organic matter also plays some role in the rates of SOD as demonstrated by the correlation between sediment δ13C and average SOD across habitats (r2 = 0.0410; p < 0.05; n = 12). The average SOD of a given habitat in turn may also play some role in determining the dark N2 fluxes versus SOD slope (Figure 2e), although this relationship is much weaker across seasons and estuaries (Figure 2f). High rates of SOD would reduce the O2 available for nitrification, and therefore reduce the efflux of N2 associated with coupled nitrification-denitrification. High rates of respiration (i.e., organic carbon is not limiting) may also allow heterotrophs to outcompete nitrifying bacteria for ammonium [Strauss and Lamberti, 2002]. In addition, labile organic carbon may allow heterotrophs to exert a stronger negative effect on nitrification when using labile rather than refractory organic carbon [Strauss and Lamberti, 2002], which may explain the low slopes at low C:N ratios (Figures 2c and 2d). In addition, high rates of respiration may result in the production of H2S during sulphate reduction which may suppress coupled nitrification-denitrification [Joye and Hollibaugh, 1995] and sulphate reducers can fix nitrogen [Nielsen et al., 2001] and thereby reduce the N2 efflux.

4.4 Light N2 Effluxes

[34] In contrast to N2 fluxes in the dark, N2 fluxes in the light were not correlated with sediment oxygen demand (r2 = 0.003). N2 fluxes were also significantly lower in the light than in the dark and were significantly different between habitats. The average ratio of dark to light N2 fluxes in the 12 habitats (i.e., how much N2 fluxes were reduced in the light) was highly correlated with net primary production (data from Maher and Eyre [2011]; Figure 5). Because average water column nitrate concentrations across most habitats were <3 µmol L−1 [Eyre and Maher, 2010], little nitrate for denitrification would have been sourced from the water column. As such, reduced N2 fluxes in the light are most likely driven by competition for ammonium and nitrate by benthic primary producers [Risgaard-Petersen et al., 1994; Sundback et al., 2000]. This competition for nitrogen exerts a stronger control on N2 fluxes in the light than does sediment oxygen demand (benthic respiration).

Figure 5.

Ratio of dark to light N2 flux versus net benthic production for each of the 12 benthic habitats. Symbol legend given in Figure 3.

4.5 Implications

[35] Detritus (nonliving organic matter) is typically considered from a perspective of its influence on food web composition and dynamics and its effect on trophic structure and biodiversity [e.g., Moore et al., 2004]. Manipulative experiments have shown that changing detrital richness may influence rates of litter decay [e.g., Moore and Fairweather, 2006] and benthic community structure [e.g., Bishop and Kelaher, 2008; Rossi et al., 2011]. More recently, it has been shown that biogeochemical processes such as benthic production and respiration, and associated nutrient fluxes, were also influenced by detrital source richness and mixes [Kelaher et al., 2013]. The quality of detritus delivered to deep-sea sediments also influences the rates of mineralization [Mayor et al., 2012]. Because denitrification permanently removes nitrogen from an ecosystem by converting fixed nitrogen to di-nitrogen gas, it is a key ecosystem process. The impact of organic matter quantity on denitrification has been well studied [e.g., Kemp et al., 1990; Caffrey et al., 1993; Eyre and Ferguson, 2009]. However, this study has demonstrated that it is not only the quantity but also the type (quality), and maybe the mixture, of organic matter that is an important control on denitrification, although we do not know the exact mechanisms involved. For example, if more labile organic matter allows heterotrophs to outcompete denitrifiers, the long-term decrease in the retention of recalcitrant particulate organic matter, as seagrasses are replaced by algae during eutrophication [McGlathery et al., 2007], may be a negative feedback with less nitrogen loss via denitrification. As such, rapid changes to detrital sources to coastal systems due to losses of mangrove, seagrasses, and saltmarshes [Mcleod et al., 2011], and associated increases in algae and macrophytes [Cloern, 2001], are also expected to impact system level losses of nitrogen via denitrification. Although SOD versus denitrification relationships have been recommended for modeling [Fennel et al., 2009], for estimating denitrification in low nitrate coastal systems [Piehler and Smyth, 2011] and upscaling to global scales [Seitzinger and Giblin, 1996; Seitzinger et al., 2006; Eyre et al., 2013], little is known about how different mixtures of organic matter influence these relationships in coastal systems. Manipulative studies looking at the effect of detrital source richness, mix, and quantity on denitrification in coastal sediments would be a fruitful area for further research.

Acknowledgments

[36] We thank M. Bautista, J. Oakes, T. Browne, and D. Erler for assistance with fieldwork. Iain Alexander is thanked for assistance with the sample and data analysis. This study was funded by the Port Macquarie Hastings Council, an Australian Postgraduate Award (APA) to D.M., and ARC Discovery (DP0342956) and ARC Linkage (LP0212073) grants awarded to B.E.

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