Eyre_et_al_GBC_2013.pdf

1 2 Quantity and quality of organic matter (detritus) drives N2 effluxes (net denitrification) across 3 seasons, benthic habitats and estuaries 4 5 6 7 8 9 Bradley D. Eyre*, Damien T. Maher, Peter Squire 10 11 Centre for Coastal Biogeochemistry, Southern Cross University, PO Box 157, Lismore, NSW 2480, 12 Australia 13 14 * Corresponding author: [email protected] 15 16 17 18 19 20 21 Running head: Organic matter driven denitrification 22 23 Abstract N2 flux rates (net denitrification) were measured over a diel cycle, seasonally, in twelve 24 benthic habitats across three warm temperate Australian coastal systems. Dark N2-N fluxes were 25 strongly controlled by sediment oxygen demand (SOD) across the three estuaries, four seasons and 26 twelve benthic habitats (r2 = 0.743; p<0.001; n=142; slope = 0.0170). However, some of the slopes 27 differed significantly between seasons, and among estuaries and habitats and all of the slopes were 28 correlated with the δ13C values and C:N ratios of sediment organic matter. Ternary mixing diagrams 29 with the contribution of algal, seagrass and terrestrial/ mangrove material to sediment organic 30 matter showed that habitats, seasons and estuaries dominated by a mixture of seagrass and algal 31 material had the lowest slopes, and slopes increase as habitats, seasons and estuaries have an 32 increasing contribution from terrestrial/ mangrove material. Overall, the slopes of dark N2 fluxes 33 versus SOD were low compared to previous studies, most likely due to either, or a combination of, 34 the C:N ratio of the organic matter, the mixture of C:N ratios making up the organic matter, the 35 structure of the organic matter, and/or the SOD rates. This study demonstrated that it is not just the 36 quantity, but also the type (quality), and maybe the mixture, of organic matter that is an important 37 control on denitrification. As such, rapid global changes to detrital sources to coastal systems due to 38 losses of mangrove, seagrasses and salt-marshes, and associated increases in algae and 39 macrophytes, are also expected to impact system level losses of nitrogen via denitrification. 40 2 41 42 1.0 Introduction Global change is rapidly altering the type, and function, of primary producers in coastal 43 ecosystems. Mangroves, seagrasses and salt-marshes have declined by around 30 to 50% over the 44 past few decades due to reclamation, deforestation and urbanisation (reviewed in Mcleod et al., 45 2011). Over the same period there has been a large increase in phytoplankton and ephemeral 46 macrophyte production (eutrophication) in many coastal systems due to nutrient over-enrichment 47 (Cloern, 2001; McGlathery et al., 2007), and an associated increase in the ratio of pelagic to benthic 48 production (Borum and Sand-Jensen, 1996; Ferguson and Eyre, 2010). The phenology of 49 phytoplankton blooms, and associated quality of phytodetritus, has been altered due to climate 50 changes (e.g. Nixon et al., 2009). Combined these changes in primary producers will significantly 51 effect the quantity, and quality, of non-living (detrital) organic matter in coastal systems, which in- 52 turn impacts ecosystem structure and function (Kelaher et al., 2013). However, it is not fully 53 understand how these changes in detrital resources influence key ecosystem processes such as 54 denitrification. 55 56 Denitrification is a critical ecosystem process that permanently removes nitrogen from an 57 ecosystem by converting fixed nitrogen to di-nitrogen gas, which can then be lost to the 58 atmosphere. Nitrogen lost to the atmosphere as di-nitrogen gas acts as a control on system level 59 primary productivity. The considered importance of denitrification is demonstrated by the 60 numerous denitrification studies in most ecosystems (Seitzinger et al., 2006). These studies have 61 identified several primary factors that control denitrification, including the supply of labile organic 62 matter and nitrate, bottom water oxygen concentrations, and several secondary factors such as the 63 presence or absence of macrofauna, macrophytes, benthic microalgae, H2S and FeS (Cornwell et 64 al., 1999; Canfield et al., 2005). In sediments where the overlying water is well-oxygenated and has 65 low nitrate concentrations, a supply of labile carbon is probably the most important controlling 66 factor on denitrification. As such, the carbon and nitrogen cycles of costal ecosystems are closely 3 67 linked via denitrification, which requires a source of organic matter (detritus) to proceed, but can 68 limit the production of organic matter via nitrogen removal. 69 70 Although many studies have looked at the effect of changes in organic matter on benthic 71 denitrification in coastal systems, these have mostly focused on organic matter quantity, not quality, 72 in response to understanding the effects of the deposition of excess organic matter production 73 (phyto-detritus) to the sediments due to nutrient over-enrichment. For example, several 74 experimental studies have shown that denitrification is enhanced when low C:N organic matter (e.g. 75 phytoplankton, glucose or yeast) is added to sediments (Brettar et al., 1992; Caffrey et al., 1993; 76 Fulweiler et al., 2008; Fulweiler et al., 2013), although coupled nitrification-denitrification, and 77 associated denitrification efficiency, can be reduced (Caffrey et al., 1993; Eyre and Ferguson, 78 2009), and Banks et al., (2013) found no response in benthic denitrification to added low C:N 79 organic matter in sediments already enriched in organic matter. Several studies have also identified 80 a positively linear relationship between dark rates of benthic denitrification and sediment oxygen 81 demand (SOD) (a proxy for the quantity of organic matter oxidation) (Table 1). Seitzinger and 82 Giblin (1996) found a slope of 0.116 for a compilation of benthic denitrification rates and SOD for 83 continental shelf sediments where the source organic matter was low C:N phyto-detritus. Similarly, 84 flow-through reactor experiments using permeable carbonate sands and seawater, with 85 phytoplankton as the organic matter source, had a SOD versus dark N2 efflux slope of 0.114 (Santos 86 et al., 2012). Organic matter oxidation would increase the supply of NH4+ from ammonification for 87 coupled nitrification–denitrification, increase the availability of electron donors for denitrification, 88 and modify the sediment redox conditions. 89 90 Much less is known about the effect of organic matter quality on benthic denitrification in 91 coastal systems. The only experimental study to compare additions of organic matter of different 92 quality on benthic denitrification rates found that high C:N organic matter added to fine muddy 4 93 sands in a temperate climate suppressed denitrification, most likely due to competition for nitrogen 94 by heterotrophs processing the refractory organic matter (Oakes et al., 2011). Similarly, Fulweiler et 95 al. (2013) speculated that the decrease in benthic denitrification during a 214 day long mesocosm 96 experiment was due to the organic matter becoming more refractory (high C:N) resulting in an 97 increased competition for nitrogen. These conditions would have been an advantage to sulphate 98 reducers that fix their own nitrogen and may also suppress nitrification, both of which would 99 decrease N2 effluxes (Fulweiler et al., 2013). Nitrate uptake in the dark (inferred to be 100 denitrification) following additions of low and high C:N organic matter to marine sediments found 101 that the magnitude of the uptake was not related to the C:N ratio, but was related to the lability 102 (structure) of the organic mater (Dahllof and Karle, 2005). 103 104 A number of previous studies in coastal systems have looked at SOD versus dark N2 efflux 105 slopes. The effect of high C:N organic matter on denitrification rates can be seen in SOD versus 106 dark N2 efflux slopes in the experiment of Oakes et al. (2011), which were 0.129 for low C:N (7.2) 107 organic matter (phyto-detritus) and 0.022 for high C:N (28.2) organic matter. A compilation of 657 108 measurements of denitrification and SOD from a range of aquatic systems found a slope of 0.086 109 (Table 1) and suggested that the lower slope than that identified by Seitzinger and Giblin (1996) 110 was simply due to the larger data set (Fennel et al., 2009), although it may have been due to the 111 mixture of organic matter (i.e. high and low C:N organic matter) across the different aquatic 112 systems. In another compilation study across multiple estuaries, seasons and habitats the average 113 slope of dark rates of denitrification versus SOD was 0.063 (Table 1), and it was suggested that the 114 lower slope compared to the earlier work of Seitzinger and Giblin (1996) was due to weaker 115 coupling between nitrification and denitrification (Piehler and Smyth, 2011), but again it may have 116 been due to the mixture of organic matter (i.e. high and low C:N organic matter) across the different 117 habitats. This same study also found a range of different slopes in different seasons and habitats 118 ranging from 0.036 to 0.107 (Table 1). Similarly, in a subtropical coastal system a low slope of 5 119 0.036 was found across multiple benthic habitats (Eyre et al., 2011a). A recent compilation of dark 120 denitrification rates and SOD measurements in permeable carbonate sands found two slopes, and it 121 was argued that the difference was due to the type of organic matter being mineralised (Eyre et al., 122 2013). A steeper slope of 0.089 was proposed to be driven by the mineralisation of episodic inputs 123 of low C:N phyto-detritus, whereas a lower slope of 0.036 was driven by inputs of high C:N organic 124 matter from coral reefs (Table 1). 125 126 Previous studies that have looked at the control of SOD on denitrification have all used rates 127 measured in the dark. Benthic denitrification rates measured in shallow coastal systems in the light 128 are usually different to dark rates due to the influence of benthic primary producers (Eyre and 129 Ferguson, 2005; Ferguson and Eyre, 2013). Benthic production during the light can increase the 130 penetration of oxygen into the sediments with contrasting effects. Coupled nitrification- 131 denitrification may be enhanced if NH4+ is readily available (Risgaard-Petersen et al., 1994; An and 132 Joye, 2001; Eyre and McKee, 2002), whereas denitrification driven by water column NO3- may 133 decrease due to consumption of NO3- by benthic microalgae and an increase in the diffusional path 134 length (Rysgaard et al., 1994). It is unknown if light denitrification rates will be well correlated with 135 SOD due to the influence of benthic primary producers. 136 137 The first hypothesis of this study is that dark rates of benthic N2 efflux (net denitrification) 138 will be strongly positively correlated with SOD, but the slope of this relationship will vary 139 depending on the sources (quality) of organic matter driving respiration (SOD). We tested this 140 hypothesis by undertaking benthic N2 efflux and SOD measurements (data in Maher and Eyre, 141 2011) over four seasons, in up to twelve benthic habitats, in three estuaries on the east coast of 142 Australia. This gave a broad range of organic matter (detritus) quality. Stable isotopes and sediment 143 molar C:N ratios were used to identify changes in the sources of organic matter driving benthic 144 respiration. The second hypothesis of this study is that SOD will not be a good predictor of light 6 145 rates of benthic denitrification due to the over-riding control of benthic production. We tested this 146 hypothesis by making the above benthic N2 efflux measurements over diel cycles. The findings 147 from this study give insight into how changes in detrital resources in coastal systems associated 148 with global change will influence key ecosystem processes such as denitrification. 149 150 2.0 Methods 151 152 153 2.1 Study Area Detailed descriptions of the three study estuaries can be found in Eyre and Maher (2010) 154 and Maher and Eyre (2011). Briefly, the estuaries are located along the South-East Australian coast 155 and fall along an estuarine maturity gradient where they have evolved by gradually infilling (Roy et 156 al., 2001). Wallis Lake is an immature stage estuary with a large central mud basin, Camden Haven 157 is an intermediate stage estuary with a more restricted shallower central basin, and the Hastings 158 River Estuary is at a mature stage with a river-dominated system characterized by river channels 159 with a highly restricted/absent central mud basin. These different stages result in distinct differences 160 in estuarine area, water residence time, catchment area and freshwater inflow (see Table 1 in Maher 161 and Eyre, 2011), the distribution of benthic habitats in each system (see maps in Eyre & Maher, 162 2010) and ecosystem scale carbon cycling (Maher and Eyre, 2012). A total of 12 benthic habitats 163 were identified in the three estuaries using a combination of underwater video, diving transects, and 164 remote sensing techniques (Table 2). The benthic habitats were delineated based on their depth 165 (subtidal, intertidal), sediment grain size (mud dominant, sand dominant), geomorphology 166 (channels, depositional basins, shoals) and dominant autotrophs (macroalgae, seagrass (Halophila 167 ovalis, Zostera capricorni, Posidonia australis and Ruppia megacarpa)). 168 169 170 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 7 171 period, although values as low as 1 mg L-1 were recorded at dawn in some seagrass areas (Maher 172 and Eyre, 2011). The estuaries have a relatively small catchment population (24 000 – 48 000; 173 Maher and Eyre, 2011) although there is modification of the catchment for agricultural uses (for full 174 description of estuarine stressors see Eyre and Maher, 2010). 175 176 177 2.2 Benthic N2 flux (net denitrification) measurements Flux sites that best represented the benthic habitat within each estuary were selected using 178 benthic habitat maps (Eyre and Maher, 2010). A total of 11, 16 and 17 sites were sampled in 179 triplicate over four seasons in the Hastings River (n=132), Camden Haven (n=192) and Wallis Lake 180 (n=204) estuaries respectively, during field campaigns in July 2006 (winter), October 2006 (spring), 181 February 2007 (summer) and April 2007 (autumn). 182 183 N2 fluxes were measured using a combination of in situ benthic chamber incubations and 184 field laboratory-based sediment core incubations. Benthic chambers were used in habitats 185 characterised by seagrass and/or large burrowing macrofauna, as the important structural elements 186 of these habitats could not be captured in cores. Cores were used in all other habitats. We believe 187 that there is no large systematic bias introduced by using a combination of cores and chambers 188 because of the high r2 between benthic fluxes measured in cores and chambers and measures which 189 are independent of the incubation device such as sediment δ13C (see Results). In addition, previous 190 studies have also shown that in shallow water coastal systems cores and chambers give comparable 191 nutrient flux data (Asmus et al. 1998). 192 193 Details of the core and chamber incubations are given in Maher and Eyre (2011). Briefly, 194 triplicate sediment cores were collected at each site with plexiglass tubes (95 mm ID x 500 mm 195 length) pushed approximately 200 mm into the sediment, leaving an overlying water volume of 196 ~1.9 L. Cores were discarded if there was any surface disturbance of sediments. Uncapped sediment 8 197 cores were placed in one of four large (150 L) perspex tanks filled with site water. Stirring was 198 maintained at a speed just below the resuspension threshold via rotating magnets throughout the 199 equilibration and incubation periods. Light was delivered by four high pressure 400 W sodium 200 bulbs (Philips Son T Agro) with in situ light conditions emulated by shading individual cores to 201 site-specific mean daily irradiance levels (± 5%). Following a 24 h equilibration period, the cores 202 were incubated over a 24 h diel cycle. 203 204 Dissolved oxygen concentrations (Hach LDO optode DO meter, ± 0.01 mg L-1) and pH and 205 temperature (Denver AP25 pH probe, ±0.001 pH units, ± 0.1°C) were measured, and alkalinity and 206 N2:Ar samples were collected at 0 (i.e., dusk) and 10 or 12 h during the dark cycle and 0 (i.e. dawn) 207 and 12 or 14 h during the light cycle. Alkalinity samples were withdrawn with a plastic syringe and 208 0.45 µm filtered water was transferred to a 10 mL milli-Q soaked and sample-rinsed polyethylene 209 vial for alkalinity analysis. To minimise the introduction of bubbles, N2 samples were collected in 210 duplicate by allowing water to flow, driven by the reservoir head, directly into 7 mL gas-tight glass- 211 stoppered glass vials filled to overflowing. The replacement water for all samples was withdrawn 212 from a sealed collapsible reservoir bag, also equilibrated at in-situ light (±5.0%) and temperature 213 (±1°C) conditions, to maintain constant Ar concentrations. Alkalinity samples were refrigerated at 214 4°C. N2:Ar samples were poisoned with 20 µl of 5% HgCl2 and stored submerged at ambient 215 temperature. N2:Ar samples were analysed on a membrane inlet mass spectrometer with oxygen 216 removal (Eyre et al., 2002). The O2, alkalinity and TCO2 (alkalinity and pH) data is presented in 217 Maher and Eyre (2011). At the conclusion of the core incubations the top 0 -10 mm depth was 218 sectioned for sediment TOC, TN and δ13C concentrations. Samples were placed in aluminium foil 219 and frozen (-20°C) until analysis. For chamber incubations, a sediment core was collected from 220 within each chamber (95 mm id, 200 mm deep) and was sampled as above. Full details of analytical 221 methods are presented in Maher and Eyre (2010). The sediment TOC and TN concentrations and 222 δ13C values are summarised in Maher & Eyre (2010) and are presented here in full. 9 223 224 Benthic chambers of a similar design to those described by Webb & Eyre (2004), but with a 225 larger volume to surface area ratio (50 l : 840 cm2), were used for in situ diel incubations in 226 macrophyte (i.e. seagrass and macroalgae) communities and shallow subtidal/intertidal habitats 227 colonised by burrowing macrofauna (predominantly the thalassinidean shrimp Trypaea 228 australiensis). Chambers were equilibrated for ≥ 24 h with the lids open, allowing free exchange of 229 water. Chambers were then sealed at dusk and sampled for O2, pH, temperature, N2: Ar and 230 alkalinity (as per core incubations) over an 18 h incubation (0h, dawn, dawn + ~ 6h) with the light 231 period capturing solar noon. 232 233 234 2.3 Calculations N2 fluxes across the sediment-water interface were calculated using the start- and end-point 235 concentration data, corrected for the addition of replacement water, as a function of incubation time, 236 core/ chamber water volume and surface area. Dark flux rates were calculated using concentration 237 data from the night-time part of incubations and light flux rates were calculated using concentration 238 data from the day-time part of incubations. Careful consideration was given to the possible effect of 239 bubbles on N2 fluxes (Eyre et al., 2002). All samples over 96% O2 saturated were excluded from N2 240 flux calculations because it was clear that these samples were affected by bubbles. As such, many 241 dark and light rates are missing due to both the initial and end dusk samples being over 96% O2 242 saturated. Bubbles have the effect of reducing N2 concentrations, giving much higher dark N2 flux 243 rates and lower light N2 flux rates or apparent rates of N-fixation. Despite the missing N2 fluxes, 244 both dark and light net denitrification rates were determined in all seasons for all estuaries, and in 245 all habitats in all estuaries, but not all habitats in all seasons in all estuaries. Because N2 fluxes were 246 measured they include both canonical denitrification and anammox. N2 fluxes reflect the balance 247 between N-fixation and denitrification and as such, are a measure of net denitrification. The terms 248 N2 efflux and net denitrification are used interchangeably. 10 249 250 251 2.4 Ternary mixing diagrams Ternary mixing diagrams were used to determine the relative contribution of algae, seagrass 252 and terrestrial/ mangrove organic matter to the sediment organic matter in each season in each 253 estuary as a whole, and in each habitat across all estuaries. However, instead of using a single end- 254 member value, a range of average measured δ13C values from the three systems and single literature 255 molar C:N (N:C) ratios were used for each end-member. N:C ratios were used in graphs for ease of 256 visualisation. The algal δ13C end-member ranged from -18.6 ± 3.8‰ (n= 73) for bulk live 257 microphytobenthos and phyto-detritus to -14.8 ± 1.2‰ for epiphytes (Maher and Eyre, 2010) with a 258 C:N ratio of 6.6 (N:C = 0.15). The bulk live seagrass end-member ranged from -15.0 ± 2.9‰ 259 (n=18) for Zostera capricorni to -11.21 ± 0.32‰ (n=6) for Posidonia australis (Maher and Eyre, 260 2010) with an C:N ratio of 20 (N:C = 0.05) (Gonneea et al., 2004). The bulk live terrestrial/ 261 mangrove end-member ranged from -32.4 ± 2.0‰ (n=18) (Maher and Eyre, 2010) to -27.0 ± 4.0‰ 262 for terrestrial organic matter (Michener and Schell, 1994; Marshall et al., 2007) with a C:N ratio of 263 50 (N:C = 0.02) (Gonneea et al., 2004). 264 265 266 267 2.5 Statistical analysis A paired t-test was used to determine if there was a statistical difference between dark and 268 light fluxes of N2 (α=0.05). One-way ANOVAs were used to explore seasonal, habitat type and 269 estuary differences for N2 fluxes (dark, light and net) and sediment TOC, TN and δ13C. Where 270 significant differences were found (i.e., p<0.05), Tukey HSD post hoc tests were used to determine 271 homogenous subsets. All analyses were done using SPSS v20.0 software. 272 273 To test for significant differences between individual sediment oxygen demand versus dark N2 274 effluxes slopes, the t statistic was calculated as a function of the difference of the slopes divided by 11 275 the difference in the standard error of the slopes, and p was calculated using the t value and n - 4 276 degrees of freedom (Kleinbaum et al 2008). 277 278 3.0 Results 279 N2 fluxes ranged from 3.5 µmol N2-N m-2 h-1 in the SS habitat in Wallis Lake in the dark 280 during autumn to 551.6 µmol N2-N m-2 h-1 in the SM habitat in Hastings River in the dark during 281 summer. N2 fluxes were significantly higher in the dark than the light (p=0.003) (Table 3). Dark N2 282 fluxes were significantly different between seasons (p=0.002), estuaries (p<0.026) and habitats 283 (p<0.001) (Table 3). N2 fluxes were higher in spring and winter than in summer and autumn. N2 284 fluxes in the Hastings River Estuary and Wallis Lakes were higher than in Camden Haven. The H, 285 IS, Z, MA, and R habitats had higher rates of net denitrification than the SS, SM, DM, P, MC, IM 286 and FMS habitats. Light N2 fluxes were only significantly different between habitats with the IS 287 and MA habitats having higher rates than the R, P, Z and IM habitats which, in turn, had higher 288 rates than the MC, H and FMS habitats. The lowest net denitrification rates in the light were in the 289 SM and SS habitats. 290 291 Sediment δ13C values ranged from -27.3 ‰ in the SS habitat during winter in the Hastings 292 River Estuary to -11.4 ‰ in the SS habitat during summer in Wallis Lakes (Table 4). Sediment δ13C 293 values were significantly different between seasons (p=0.02), estuaries (p<0.001) and habitats 294 (p<0.001) (Table 4). In summer and spring sediment δ13C values were more enriched than in winter 295 and autumn. Sediment organic matter in Wallis Lake had more enriched δ13C values than in 296 Camden Haven, which in turn, had more enriched δ13C values than in the Hastings River Estuary. 297 The MA habitat had the most enriched sediment δ13C values, followed by the P and R habitats, then 298 the Z and H habitats, and then the SM, DM, IM, IS, SS and MC habitats. The FMS habitat had the 299 most depleted sediment δ13C values. 300 12 301 Sediment total organic carbon (TOC) concentrations ranged from 0.07% in the MC habitat 302 during summer and autumn in Wallis Lake to 7.51% in the FMS habitat during winter in the 303 Hasting River Estuary (Table 4). TOC concentrations were significantly different between estuaries 304 (p<0.001) and habitats (p<0.001), but not seasons (p=0.979). Sediments in Wallis Lake had the 305 highest TOC concentrations, followed by Camden Haven, with the lowest sediment TOC 306 concentrations in the Hasting River Estuary. The highest TOC concentrations were in the R and 307 MA habitats, followed by the SM, DM, Z, H and FMS habitats and then the P, MC and IM habitats. 308 The SS and IS habitats had the lowest TOC concentrations. 309 310 The patterns of sediment nitrogen concentrations were similar to those for TOC (Table 4). 311 Sediment nitrogen concentrations ranged from 0.02% in a number of habitats during different 312 seasons in all systems, to 0.54% in the Z habitat during spring in Wallis Lake. Sediment nitrogen 313 concentrations were significantly different between seasons (p<0.001), estuaries (p<0.001) and 314 habitats (p<0.001). Sediments in Wallis Lake had the highest nitrogen concentrations, followed by 315 Camden Haven, with the lowest sediment nitrogen concentrations in the Hasting River Estuary. The 316 highest sediment nitrogen concentrations were in the R and MA habitats, followed by the SM, DM, 317 Z, H and FMS habitats and then the P, MC and IM habitats. The SS and IS habitats had the lowest 318 sediment nitrogen concentrations. 319 320 4.0 Discussion 321 4.1 Comparison of Denitrification Rates 322 Few studies have looked at denitrification rates across multiple habitats within a coastal 323 system (e.g. Piehler and Smyth, 2011), and only one of these has measured denitrification rates over 324 a diel cycle (Eyre et al., 2011a). Subtidal, intertidal and seagrass habitats are the only common 325 habitats across all these multiple habitat studies. Similar to this study, most multiple habitat studies 326 have found higher dark rates of denitrification in seagrass habitats compared to intertidal and 13 327 subtidal shoals, although the higher rates may only occur in some seasons (Eyre et al, 2011a; Eyre 328 et al., 2011b; Piehler and Smyth, 2011; Smyth et al., 2012). These more recent high rates of net 329 denitrification using the N2: Ar technique (28 to 824 µmol m-2 h-1) contrast with earlier very low 330 rates (<1 to 35 µmol m-2 h-1) of denitrification measured in seagrass communities using isotope 331 pairing (e.g. Risgaard-Petersen et al., 1998; Risgaard-Petersen and Ottosen, 2000; Welsh et al. 332 2000). Eyre et al., (2011a) suggested that such differences may reflect a difference between tropical 333 and temperate systems, with higher rates in tropical systems and lower rates in temperate systems. 334 However, Piehler and Smyth (2011) and Smyth et al. (2012) recently measured high rates of net 335 denitrification in temperate seagrass communities using N2: Ar ratios. This suggests that the 336 differences are either methodological, with lower rates measured using isotope pairing and higher 337 rates measured using N2: Ar, or due to the biogeochemistry of the different seagrass species. For 338 example, low rates of coupled nitrification-denitrification might reflect irregular oxygen release by 339 the roots of some temperate seagrass species (Frederiksen and Glud, 2006). Further work using both 340 methods simultaneously in different seagrass communities is required. 341 342 343 4.2 Organic matter control on N2 effluxes Dark N2 fluxes were strongly controlled by SOD (data from Maher and Eyre, 2011) across 344 the three estuaries, four seasons and twelve benthic habitats (r2 = 0.743; p<0.001; n=142; slope = 345 0.0170; Fig. 1; Table 5). Dark N2 fluxes also showed a strong correlation with dark TCO2 effluxes 346 (r2 = 0.499; p<0.0001; n=142; slope= 0.0147; TCO2 data from Maher and Eyre, 2011) 347 demonstrating that much of the SOD is due to organic matter oxidation. Average water column 348 nitrate concentrations across most habitats were <3 µmol L-1 (Eyre and Maher, 2010), suggesting 349 that much of the N2 flux was driven by coupled nitrification-denitrification. As such, the control of 350 SOD on N2 effluxes most likely reflects an increased supply of NH4+ from ammonification for 351 coupled nitrification–denitrification, the availability of electron donors for denitrification, and 352 modification of sediment redox conditions. 14 353 354 There were differences however, in the slope of dark N2 fluxes versus SOD (benthic 355 respiration) between seasons, between estuaries, and between habitats (Table 5). Although only 356 some of the slopes were significantly different from each other (Tables 6 and 7), all the different 357 slopes were highly correlated with δ13C values of sediment organic carbon across the different 358 habitats (Fig. 2a), and different seasons, in different estuaries (Fig. 2b). The δ13C value of organic 359 matter reflects its source (Peterson, 1999) and, as such, the δ13C values of sediment organic carbon 360 reflect the mixture of different sources of organic matter within the sediments (see later discussion 361 of changes during mineralisation). This suggests that the slope of dark N2 fluxes versus benthic 362 respiration is driven by the type, or mixture, of organic matter undergoing mineralization (Eyre et 363 al., 2013). An increase in the gradient of the slope was related to more depleted sediment δ13C 364 values. Sediment C:N ratios were also correlated with the slope of dark N2 fluxes versus benthic 365 respiration across both the different habitats (Fig. 2c) and different seasons in different estuaries 366 (Fig. 2d). Interestingly, higher C:N ratios were related to an increase in the gradient of the slope. 367 However, the C:N relationship was weaker than with sediment δ13C values, demonstrating that the 368 type, or mixture, of organic matter undergoing mineralization was a more important control than its 369 C:N ratio. 370 371 Ternary mixing diagrams showing the contributions of algal, seagrass and terrestrial/ 372 mangrove material to sediment organic matter give some insight into how changes in the type, or 373 mixture, of organic matter influence the slope of dark N2 fluxes versus benthic respiration across 374 different habitats, estuaries and seasons (Figs. 3 and 4). Sediments dominated by approximately 375 equal mixtures of seagrass and algal material have the lowest slopes. The slopes increase for 376 habitats that also have a contribution from terrestrial/ mangrove material, with the steepest slope in 377 the FMS habitat reflecting approximately equal proportions of terrestrial/ mangrove and algal 378 material. Similarly, the lowest slopes were in Wallis Lake where organic matter in the sediments 15 379 was dominated by equal mixtures of algae and seagrass, and the highest slopes were in the Hastings 380 River Estuary where terrestrial/ mangrove material made a larger contribution (Fig. 4). The slopes 381 also increased in the seasons that had the largest contribution of terrestrial/ mangrove material, such 382 as winter in the Hastings River Estuary, summer in Camden Haven and autumn in Wallis Lake (Fig. 383 4). These seasonal interactions probably reflect differences in the delivery and trapping of terrestrial 384 organic material due to different estuarine geomorphology. 385 386 4.3 Processes driving organic matter control on N2 effluxes 387 Although it is quite clear that the type, or mixture, of organic matter is driving the dark N2 388 fluxes versus SOD slopes across habitats, seasons and estuaries (Figs. 2a and 2b), the mechanisms 389 involved were not entirely clear. It may be due to one, or a combination of, the C:N ratio of the 390 organic matter, the mixture of C:N ratios making up the organic matter, the structure of the organic 391 matter, and/or the overall SOD rates. All this mechanisms would influence the community 392 composition of total bacteria, ammonia-oxidizing archaea and denitrifiers, which have also been 393 correlated to sediment δ13C in an estuarine system (Abell et al., 2013). We will explore each of 394 these mechanisms below. 395 396 Overall the dark N2 fluxes versus SOD slopes for this study were low (0.0114 to 0.0332; 397 Table 5) compared to most previous compilation studies (Table 1). For example, a slope of 0.116 398 has been reported for continental shelf sediments (Seitzinger and Giblin, 1996), a slope of 0.086 399 was found for a compilation of 657 denitrification and SOD measurements (Fennel et al., 2009), 400 and an estuarine study across multiple seasons and habitats found an average slope of 0.063 401 (Piehler and Smyth, 2011) (Table 1). Multiple lines of evidence suggest that higher slopes (e.g. 402 Seitzinger and Giblin, 1996) are driven by the decomposition of low C:N phyto-detritus and lower 403 slopes (this study) may be driven by the decomposition of mostly high C:N sources of organic 404 matter such as seagrass, mangrove and terrestrial material. The lower slopes in Fennel et al. (2009) 16 405 and Piehler and Smyth (2011) are probably also due to a mixture of organic matter sources driving 406 denitrification in these studies, whereas phyto-detritus would be the dominant source of organic 407 matter to the continental shelf sediments that had the higher slope of 0.116 (Seitzinger and Giblin, 408 1996). Flow-through reactor experiments using permeable carbonate sands and seawater, with 409 phytoplankton as the organic matter source, had a benthic respiration versus dark N2 efflux slope of 410 0.114 (Santos et al., 2012). The only experimental study to compare additions of organic matter of 411 different quality on dark N2 fluxes found a slope of 0.129 for low C:N (7.2) organic matter (phyto- 412 detritus) and a slope of 0.022 for high C:N (28.2) organic matter (slopes calculated from Figure 1 in 413 Oakes et al., 2011). Eyre et al. (2013) also speculated that high C:N organic matter from coral reefs 414 may be driving low dark N2 fluxes versus benthic respiration slopes (0.036) in permeable carbonate 415 sediment with higher slopes (0.089) driven by episodic input of low C:N phyto-detritus. 416 417 The dark N2 fluxes versus SOD slopes may also be driven by high C:N organic matter 418 which would release less nitrogen as N2 for a given amount of respiration. A lack of N release 419 would be enhanced due to competition for nitrogen by heterotrophic bacteria (Oakes et al., 2011); 420 N-limitation of the microbial decomposition of high C:N organic material results in the uptake and 421 accumulation of nitrogen by heterotrophic bacteria (Tupas and Koike, 1991; van Duyl et al., 1993; 422 Lomstein et al., 1998). Several types of bacteria can assimilate NH4+, including sulphate reducers 423 and fermentative bacteria (Koike and Sumi 1989). Additionally, sulphate reducers can fix nitrogen 424 (Nielsen et al., 2001). N2 effluxes are a measure of denitrification minus N-fixation and as such, 425 increased N-fixation will result in a reduced N2 efflux. Further, coupled nitrification-denitrification 426 may be suppressed by H2S produced during sulphate reduction (Joye and Hollibaugh, 1995; 427 Fulweiler et al., 2013). Consistent with higher rates of heterotrophic N-fixation during 428 decomposition of higher C:N organic material is the reduction in dark N-fixation rates in permeable 429 carbonate sands when low C:N phyto-detritus was deposited (Eyre et al., 2008). Nitrogen 430 assimilation and N-fixation by heterotrophic bacteria would be a nitrogen conservation process in 17 431 high C:N seagrass communities that would otherwise lose large amounts of nitrogen via 432 denitrification due to high rates of benthic respiration, i.e. the loss of nitrogen via denitrification per 433 unit of respiration is an order of magnitude lower (Table 5) than the loss in a low C:N environment 434 such as the continental shelf (Seitzinger and Giblin, 1996)(Table 1). 435 436 If the C:N ratios were an important control on the SOD versus N2 efflux relationships in this 437 study, it would be expected that higher slopes would be driven by low C:N algal material. However, 438 the steeper slopes occur in sediments with higher amounts of high C:N terrestrial/ mangrove 439 organic matter, which is the opposite of what was expected. Because of the stronger relationship 440 between the slopes and the δ13C than compared with C:N ratios, one explanation may be that the 441 structure of the decomposing organic matter is influencing the slope (Fig. 2 a-d). For example, 442 Dahllof and Karle (2005) found that the C:N ratio of organic matter added to sediments was not an 443 important control on nitrate uptake in the dark (inferred to be due to denitrification). Instead they 444 argued that nitrate uptake was related to the structure of the organic carbon, with greater nitrate 445 uptake driven by organic matter containing easily degraded lipids and starch, and lower nitrate 446 uptake driven by organic matter containing relatively refractory cellulose and lignin. This would 447 explain the low dark N2 fluxes versus SOD slopes in the seagrass habitats and Wallis Lake, which is 448 dominated by seagrass production (Maher and Eyre, 2011), but not the high slopes in the habitats 449 and systems with higher inputs of terrestrial/ mangrove organic matter. 450 451 Alternatively, the dark N2 fluxes versus SOD slopes in this study may reflect different components 452 of the high C:N organic matter or a mixture of low and high C:N organic matter. High C:N organic 453 matter dominates within the study systems, giving the low slopes compared to previous studies, but 454 the proportion of the labile component of the high C:N organic matter, or the low C:N proportion of 455 the organic matter mixture, driving the dark N2 fluxes versus benthic respiration slopes may vary 456 across the habitats, seasons and estuaries. For example, slopes increase for habitats with increasing 18 457 amounts of terrestrial/ mangrove material, but the SS and IS habitats that have higher proportions of 458 algal material for a given δ13C values also have slightly higher slopes (Figure 3). Even more subtle 459 is the P habitat, which has a slightly higher algal contribution and associated higher slope than the R 460 and H seagrass communities for a given δ13C value (Figure 3). The sediment organic matter δ13C 461 value and molar C:N ratio is a measure of the remaining organic matter after the more labile organic 462 matter has been mineralized. This is consistent with more depleted sediment δ13C values associated 463 with higher slopes, as removal of more labile 13C-enriched cellulose material results in more 13C 464 depleted lignin-rich material remaining in the sediments (Benner et al., 1987). 465 466 The type, or mixture, of organic matter also plays some role in the rates of SOD as 467 demonstrated by the correlation between sediment δ13C and average SOD across habitats 468 (r2=0.0410; p<0.05; n=12). The average SOD of a given habitat in turn may also play some role in 469 determining the dark N2 fluxes versus SOD slope (Figure. 2e), although this relationship is much 470 weaker across seasons and estuaries (Figure 2f). High rates of SOD would reduce the O2 available 471 for nitrification, and therefore reduce the efflux of N2 associated with coupled nitrification- 472 denitrification. High rates of respiration (i.e. organic carbon is not limiting) may also allow 473 heterotrophs to outcompete nitrifying bacteria for ammonium (Strauss and Lamberti, 2002). In 474 addition, labile organic carbon may allow heterotrophs to exert a stronger negative effect on 475 nitrification when using labile rather than than refractory organic carbon (Strauss and Lamberti, 476 2002), which may explain the low slopes at low C:N ratios (Figures 2c and 2d). In addition, high 477 rates of respiration may result in the production of H2S during sulphate reduction which may 478 supress coupled nitrification-denitrification (Joye and Hollibaugh, 1995) and sulphate reducers can 479 fix nitrogen (Nielsen et al., 2001) and thereby reduce the N2 efflux. 480 481 482 4.4 Light N2 effluxes In contrast to N2 fluxes in the dark, N2 fluxes in the light were not correlated with sediment 19 483 oxygen demand (r2=0.003). N2 fluxes were also significantly lower in the light than in the dark and 484 were significantly different between habitats. The average ratio of dark to light N2 fluxes in the 485 twelve habitats (i.e., how much N2 fluxes were reduced in the light) was highly correlated with net 486 primary production (data from Maher and Eyre, 2011) (Figure 5). Because average water column 487 nitrate concentrations across most habitats were <3 µmol L-1 (Eyre and Maher, 2010), little nitrate 488 for denitrification would have been sourced from the water column. As such, reduced N2 fluxes in 489 the light are most likely driven by competition for ammonium and nitrate by benthic primary 490 producers (Risgaard et al., 1994; Sundback et al., 2000). This competition for nitrogen exerts a 491 stronger control on N2 fluxes in the light than does sediment oxygen demand (benthic respiration). 492 493 494 4.5 Implications Detritus (non-living organic matter) is typically considered from a perspective of its 495 influence on food web composition and dynamics and its effect on trophic structure and 496 biodiversity (e.g., Moore et al., 2004). Manipulative experiments have shown that changing detrital 497 richness may influence rates of litter decay (e.g., Moore and Fairweather, 2006) and benthic 498 community structure (e.g. Bishop and Kelaher, 2008; Rossi et al., 2011). More recently, it has been 499 shown that biogeochemical processes such as benthic production and respiration, and associated 500 nutrient fluxes, were also influenced by detrital source richness and mixes (Kelaher et al., 2013). 501 The quality of detritus delivered to deep-sea sediments also influence rates of mineralisation 502 (Mayor et al., 2012). Because denitrification permanently removes nitrogen from an ecosystem by 503 converting fixed nitrogen to di-nitrogen gas it is a key ecosystem process. The impact of organic 504 matter quantity on denitrification has been well studied (e.g., Kemp et al., 1990; Caffrey et al., 505 1993; Eyre and Ferguson, 2009). However, this study has demonstrated that it is not just the 506 quantity but also the type (quality), and maybe the mixture, of organic matter that is an important 507 control on denitrification, although we do not know the exact mechanisms involved. For example, if 508 more labile organic matter allows heterotrophs to outcompete denitrifiers, the long-term decrease in 20 509 the retention of recalcitrant particulate organic matter as seagrasses are replaced by algae during 510 eutrophication (McGlathery et al., 2007), may be a negative feedback with less nitrogen loss via 511 denitrification. As such rapid changes to detrital sources to coastal systems due to losses of 512 mangrove, seagrasses and salt-marshes (Mcleod et al., 2011), and associated increases in algae and 513 macrophytes (Cloern, 2001), are also expected to impact system level losses of nitrogen via 514 denitrification. Although SOD versus denitrification relationships have been recommended for 515 modeling (Fennel et al., 2009), for estimating denitrification in low nitrate coastal systems (Piehler 516 and Smyth, 2011) and up-scaling to global scales (Seitzinger and Giblin, 1996; Seitzinger et al., 517 2006; Eyre et al., 2013), little is known about how different mixtures of organic matter influence 518 these relationships in coastal systems. Manipulative studies looking at the effect of detrital source 519 richness, mix and quantity on denitrification in coastal sediments would be a fruitful area for further 520 research. 521 522 Acknowledgements 523 We thank M. Bautista, J. Oakes, T. Browne, and D. Erler for assistance with fieldwork. Iain 524 Alexander is thanked for assistance with the sample and data analysis. This study was funded by 525 Port Macquarie Hastings Council, an Australian Postgraduate Award (APA) to D.M. and ARC 526 Discovery (DP0342956) and ARC Linkage (LP0212073) grants awarded to B.E. 527 528 References 529 Abell, G.C.J., Ross, D.J., Keane, J.P., Oakes, J.M., Eyre, B.D., Robert, S.S., and Volkman, J.K. 530 (2013), Nitrifying and denitrifying microbial communities and their relationship to nutrient fluxes 531 and sediment geochemistry in the Derwent Estuary, Tasmania. Aquatic Microbial Ecology 70, 63- 532 75. 533 21 534 An, S., and S. B. Joye (2001), Enhancement of coupled nitrification-denitrification by benthic 535 photosynthesis in shallow estuarine sediments, Limnol. Oceanogr., 46, 62-74. 536 537 Asmus R (1986), Nutrient flux in short-term enclosures of intertidal sand communities. OPHELIA 538 26, 1-18. 539 540 Banks, J. L., Ross, D. J., Keough, M. 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Eyre (2004), The effects of two benthic chamber stirring systems: a 764 conventional rotor and a two dimensional flow system, over the diffusive boundary layer, solute 765 efflux and passive flow through a macrofauna borrow. Estuaries, 27, 353-362. 766 767 Welsh D. T., M. Bartoli, D. Nizzoli, G, Castaldelli, S.A. Riou, P. Viaroli (2000), Denitrification, 768 nitrogen fixation, community primary production and inorganic-N and oxygen fluxes in an 769 intertidal Zostera noltii meadow. Mar. Ecol. Prog. Ser., 208, 65-77 770 771 772 773 774 31 775 Figure captions. 776 777 Fig 1. Dark N2 flux (µmol N2-N m-2 h-1) versus sediment oxygen demand (µmol O2 m-2 h-1) across 778 the three estuaries, four seasons, and twelve benthic habitats. 779 780 Fig. 2. Sediment δ13C versus the slope of dark N2 flux (µmol N2-N m-2 h-1) versus sediment oxygen 781 demand (µmol m-2 h-1) (a) across the twelve benthic habitats, and (b) for each season in each 782 estuary; molar C:N ratio versus the slope of dark N2 flux (µmol N2-N m-2 h-1) versus sediment 783 oxygen demand (µmol O2 m-2 h-1) (c) across the twelve benthic habitats, (d) for each season in each 784 estuary; and average sediment oxygen demand versus the slope of dark N2 flux (µmol N2-N m-2 h- 785 1 786 each season in each estuary. 787 788 789 Fig. 3. Ternary mixing diagram showing the relative contribution of algae, seagrass and terrestrial/ 790 mangrove organic matter to each habitat. Also shown for each habitat is the slope of dark N2-N flux 791 versus sediment oxygen demand (from Table 5). Habitat abbreviations given in Table 2. N:C ratios 792 were used for ease of visualisation. C:N ratio of 6.6 = N:C ratio of 0.15. C:N ratio of 20 = N:C ratio 793 of 0.05. C:N ratio of 50 = N:C ratio of 0.02. ) versus sediment oxygen demand (µmol O2 m-2 h-1) (e) across the twelve benthic habitats, (f) for 794 795 Fig. 4. Ternary mixing diagram showing the relative contribution of algae, seagrass and terrestrial/ 796 mangrove organic matter in the three estuaries for each season. Also shown for each estuary for 797 each season is the slope of dark N2-N flux versus sediment oxygen demand (from Table 5). C:N 798 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 799 0.02. 800 801 Fig. 5. Ratio of dark to light N2 flux versus net benthic production for each of the twelve benthic 32 802 habitats. Symbol legend given in Figure 3. 803 33 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). All regressions were forced through zero as there would be no N2 efflux when respiration equals zero. Sediment Type Slope Source Subtropical to temperate continental shelf muds 0.116 Seitzinger and Giblin 1996 Subtropical to temperate muds and sands 0.086 Fennel et al., 2009 Very coarse permeable carbonate sands 1 Eyre et al., 2008 Temperate muds, range of habitats from subtidal 0.036 to 0.107 shoals to seagrass 0.033 Piehler and Smyth, 2011 0.063 (average) 2 0.036 Eyre et al., 2011a Fine muddy sand, low C:N organic matter 3 0.129 Oakes et al., 2011 Fine muddy sand, high C:N organic matter 3 0.022 Oakes et al., 2011 Very coarse permeable carbonate sands, 0.114 Santos et al., 2012 0.019 to 0.047 Eyre et al., 2013 Subtropical muds, range of habitats from subtidal shoals to seagrass low C:N organic matter Fine to very coarse permeable carbonate sands, high C:N organic matter Fine to very coarse permeable carbonate sands, 0.035 (average) 0.089 Eyre et al., 2013 low C:N organic matter 804 1 Data from Fig. 4 in Glud et al. (2008) and Fig. 3 in Eyre et al. (2008); 2Data from Fig. 5 in Eyre et 805 al., 2011a, r2 is slightly less than in Eyre et al., 2011a due to forcing through zero. 3Slopes 806 calculated from Figure 1 in Oakes et al. (2011). 34 807 Table 2. Benthic habitats, descriptions and abbreviations (modified from Maher and Eyre, 2010). Habitat Description Abbreviation Intertidal sand Intertidal habitats composed primarily of sand IS Intertidal mud Intertidal habitats composed primarily of mud/ silt IM Subtidal sands Subtidal shoal habitats, composed of sand SS Subtidal muds Subtidal shoal habitats, composed of mud/silt SM Fluvial muds and sands Channel habitats in the mid/upper estuary dominated FMS by fluvial sands and/or muds Marine channel Channel habitats, lower estuary, clean quartz sands MC Deep subtidal muds Subtidal habitats >4 m deep dominated by muds DM Zostera Seagrass habitats with Zostera capricorni dominant Z Halophila Seagrass habitats with Halophila australis dominant H Ruppia Seagrass habitats with Ruppia megacarpa dominant R Macroalgae Habitats dominated by macroalgae MA Posidonia Seagrass habitats with Posidonia australis dominant P 808 35 809 810 Table 3. Seasonal dark and light N2 fluxes (µmol N2-N m-2 h-1) in the twelve habitats across Wallis Lake, Camden Haven and the Hastings River Estuary (mean ± SE, n=2 or 3). ND = many dark and light rates are missing due to both the initial and end dusk samples being over 96% O2 saturated. Habitat Winter Spring Summer Autumn Dark Light Dark Light Dark Light Dark Light FMS ND ND 17.8 17.8 25.8 ± 8.6 47.3 ± 10.3 98.3 ± 25.5 65 ± 16.3 H ND ND ND ND 65.3 ND 19.1 ± 18.5 13.1 ± 5.5 IM ND 100.2 ± 13.7 48.7 ± 0.3 ND 89.8 ± 65.6 ND 35.6 ND IS 89.9 ND 172.3 ± 118.5 157.6 ± 45.6 212.6 ± 9.3 173.7 ± 29.3 234.2 257 ± MC ND 74 ± 9.3 ND ND 25.1 13.1 ± 8 42.3 59.8 ± 7.5 SM 18.9 ± 12.9 ND ND ND ND 5.5 ± 2.1 8.4 ± 2.9 ND SS ND ND ND ND 16.2 ± 3.1 28.4 ± 11 13.5 ± <0.1 21 ± 11.4 Z ND ND 279.9 ND 445.2 ± 23.4 ND 247.5 ± 48.9 ND FMS ND 13.2 5.3 ± 3.1 8.2 58.9 ± 17.9 53 ± 32.8 16.3 2.4 H ND ND ND ND ND 12.3 ND 21.4 IM ND ND ND 83.3 ± 15.3 227.4 ND 75.0 189.6 IS ND ND 11.1 ± 7.9 ND 40.3 ND ND ND M IS ND ND 87.9 ± 20.7 17 ± 9.8 ND ND ND ND MZ ND ND 118.8 ND ND ND ND ND MC ND ND 26.2 21.2 ± 15.8 20.0 ± 5.2 45.6 ± 38.1 9.3 ± 4.1 13.8 Hastings River Camden Haven R ND ND ND ND 206.1 29.2 ± 15.5 86.6 ± 1.8 147.8 ± 38.4 SM 32.9 ± 6.5 26.5 ± 7.5 31.4 ± 4.0 13.8 ± 4.2 44.9 ± 10.2 25.7 ± 10.2 4.9 7.8 ± 4.7 SS ND ND 30.7 4.7 ± 0.3 13.7 18.3 14.5 18 Z 135.5 ± 24.2 ND ND 118.1 ND 75.8 102.7 ± 15.1 ND Wallis Lake DM ND ND 17.2 8.6 ± 3.9 ND 11.6 ± 5.9 ND 15.6 ± 5.8 FMS ND 31.8 11.5 ± 5.2 8.1 ± 3.1 8.8 ± 0.1 37.3 ± 13.8 53.5 ± 4.6 ND H ND ND 62 ND 121.8 ± 7.2 75.1 ± 28.2 119.2 ± 32.0 ND IM ND ND 11.9 ± 8.0 61.8 11.4 ND 21.7 ± 13.3 10.9 IS ND ND 19.9 ND 6.7 ± 3.7 ND 12.7 ± 6.2 ND MA 34.7 ND ND ND ND 205.6 ± 66.7 199.4 ND MC ND ND 0.1 12.2 12.6 ± 1.9 43.5 ND ND P 78 ± 40.4 55.1 ND 93.8 ± 12.3 50.6 ND 188.3 33 R ND ND ND ND 244.7 ± 38.2 68.1 ± 38.9 132.1 ± 24.8 281.6 ± 57.9 SM ND ND 7.2 8.3 12.9 ± 1.1 4.3 13.6 ND SS ND ND ND ND ND 7.5 3.5 ND Z ND ND ND ND 150.9 ± 24.1 ND 203.6 ± 61 78.4 811 812 813 37 814 815 Table 4. Seasonal sediment δ13C, total organic carbon and total nitrogen concentrations (%) in the twelve habitats across Wallis Lake, Camden Haven and the Hastings River Estuary. Habitat abbreviations given in Table 2. Habitat δ13C Winter TOC TN δ13C Spring TOC TN δ13C Summer TOC δ13C TN Autumn TOC TN Hastings River FMS -25.61 ± 0.51 H ND IM -24.23 ± 0.6 IS -24.42 ± 1.02 MC -22.91 ± 0.57 SM -26.88 ± 0.83 SS -27.28 ± 0.7 0.83 ± 0.21 ND 1.07 ± 0.15 0.29 ± 0.09 3.67 ± 0.04 0.4 ± 0.13 0.21 ± 0.04 0.06 ± 0.01 ND 0.07 ± 0.01 0.02 ± 0.01 0.24 0.02 ± 0.01 0.02 ± 0.01 -24.14 ± 0.32 ND -23.24 ± 0.49 -18.76 ± 0.32 -23.87 ± 0.11 -24.15 ± 0.24 -20.91 ± 0.57 0.48 ± 0.12 ND 2.28 ± 1.35 0.22 ± 0.04 4.88 ± 0.65 0.72 ± 0.12 0.2 ± 0.03 0.05 ± 0.01 0.1 0.15 ± 0.06 0.03 ± 0.01 0.25 0.05 ± 0.01 0.02 -22.83 ± 0.77 ND -21.52 ± 1.24 -19.08 ± 0.18 -19.98 ± 0.4 -21.92 ± 0.31 -24.07 ± 0.85 0.26 ± 0.09 1.95 ± 0.14 0.79 ± 0.11 0.45 ± 0.15 3.36 ± 0.02 0.58 ± 0.07 0.2 ± 0.03 0.05 ± 0.01 0.16 ± 0.01 0.08 ± 0.01 0.04 ± 0.01 0.26 0.05 ± 0.01 0.02 -23.38 ± 0.34 ND -22.46 ± 0.1 -18.99 ± 0.32 -23.95 ± 0.04 -22.72 ± 0.1 -21.84 ± 0.29 0.44 ± 0.09 2.01 ± 0.18 0.75 ± 0.12 0.46 ± 0.19 3.06 ± 0.06 1.86 ± 0.25 0.23 ± 0.04 0.05 ± 0.01 0.13 ± 0.01 0.06 ± 0.01 0.04 ± 0.02 0.2 0.12 ± 0.01 0.03 ± 0.01 Z -20.57 ± 0.56 1.01 ± 0.19 0.07 ± 0.01 -19.72 ± 0.33 0.77 ± 0.24 0.07 ± 0.02 -19.74 ± 0.07 0.85 ± 0.05 0.08 ± 0.01 ND ND ND Camden Haven FMS -23.4 ± 0.28 H ND IM -15.43 ± 0.06 IS -21.62 ± 0.18 M IS -22.26 ± 0.74 MZ -21.02 ± 0.32 MC -21.93 ± 0.67 R -15.81 ± 0.07 SM -20.88 ± 0.72 SS -24.83 ± 0.92 Z -16.33 ± 0.05 7.51 ± 3.22 4.07 ± 0.1 3.08 ± 0.17 0.42 ± 0.06 0.42 ± 0.05 1.05 ± 0.15 0.11 ± 0.01 2.81 ± 0.52 2.83 ± 0.32 0.1 ± 0.01 4.04 ± 0.05 0.46 ± 0.14 0.38 ± 0.01 0.3 ± 0.02 0.04 0.03 0.07 ± 0.01 0.01 0.28 ± 0.05 0.21 ± 0.03 0.02 0.37 -23.81 ± 0.65 ND -15.79 ± 0.16 -21.29 ± 0.24 -20.27 ± 0.72 -20.03 ± 0.18 -21.44 ± 0.62 -15.84 ± 0.56 -20.77 ± 0.46 -21.14 ± 0.87 -15.87 ± 0.17 2.9 ± 0.52 3.92 ± 0.19 1.61 ± 0.4 0.24 0.82 ± 0.19 2.76 ± 1.08 0.15 ± 0.03 2.42 ± 0.91 2.07 ± 0.23 0.13 ± 0.01 4.44 ± 0.05 0.16 ± 0.04 0.35 ± 0.02 0.17 ± 0.04 0.03 0.06 ± 0.01 0.2 ± 0.08 0.02 0.24 ± 0.08 0.15 ± 0.02 0.02 0.42 -21.67 ± 0.49 ND ND ND ND ND -21.21 ND -21.01 -20.76 ± 0.08 ND 2.16 ± 0.75 3.01 ± 0.11 2.02 ± 0.74 0.5 ± 0.09 0.64 ± 0.02 2.42 ± 0.4 0.12 ± 0.01 3.35 ± 0.18 1.94 ± 0.21 0.17 ± 0.03 4.04 ± 0.04 0.12 ± 0.03 0.27 ± 0.01 0.2 ± 0.07 0.05 ± 0.01 0.06 0.18 ± 0.03 0.02 0.34 ± 0.01 0.19 ± 0.01 ND 0.37 ± 0.01 -23.21 ± 0.48 -15.81 ± 0.11 -17.65 ± 0.56 -21.42 ± 0.11 -20.46 ± 0.27 -17.58 ± 0.28 -16 ± 0.04 -20.31 ± 0.63 ND -16.4 ± 0.34 4.62 ± 0.77 4.28 ± 0.08 2.35 ± 0.23 0.69 ± 0.04 2.17 ± 0.35 1.82 ± 0.14 0.13 1.64 ± 0.43 2.49 ± 0.13 ND 3.78 ± 0.47 0.23 ± 0.03 0.32 ± 0.03 0.23 ± 0.02 0.08 ± 0.01 0.15 ± 0.02 0.14 ± 0.01 0.02 0.15 ± 0.04 0.2 ± 0.01 ND 0.28 ± 0.03 Wallis Lake DM -16.72 ± 1.09 FMS -21.45 ± 0.32 H ND IM -20.13 ± 1.32 IS -19.79 ± 1.18 MA -13.85 ± 0.05 MC -23.59 ± 1.36 P -14.83 ± 0.7 R -13.95 ± 0.05 SM ND SS -15.97 ± 0.53 Z -16.94 ± 0.38 3.64 ± 0.35 3.49 ± 0.17 3.76 ± 0.3 1.12 ± 0.27 0.49 ± 0.17 3.55 ± 0.18 ND 1.03 ± 0.08 3.73 ± 0.03 ND 0.3 ± 0.05 1.02 ± 0.21 0.31 ± 0.03 0.24 ± 0.02 0.35 ± 0.03 0.09 ± 0.02 0.05 ± 0.01 0.33 ± 0.01 ND 0.09 ± 0.01 0.35 0.12 ± 0.01 0.03 0.08 ± 0.02 -16.44 ± 0.69 -22.77 ± 0.11 -13.98 -21.61 ± 0.02 -20.05 ± 0.58 -13.7 ± 0.07 ND -13.86 ± 0.17 -14.31 ± 0.25 -22.25 ± 0.15 -13.37 ± 0.18 -14.73 ± 0.05 3.4 ± 0.35 3.82 ± 0.27 2.96 ± 0.8 1.98 ± 0.13 0.7 ± 0.02 2.97 ± 0.6 ND 1.13 ± 0.21 4.08 ± 0.38 2.28 ± 0.11 0.41 ± 0.02 5.59 ± 0.38 0.33 ± 0.04 0.26 ± 0.02 0.29 ± 0.08 0.17 ± 0.01 0.07 0.29 ± 0.04 ND 0.11 ± 0.02 0.39 ± 0.04 0.17 ± 0.01 0.05 ± 0.01 0.54 ± 0.04 -16.62 ± 0.66 -22.56 ± 0.69 -13.66 ± 0.04 -21.25 ± 0.1 -18.03 ± 0.59 -13.85 ± 0.1 -19.13 ± 0.5 -14.73 ± 0.59 -14.19 ± 0.08 -22.17 ± 0.11 ND -14.35 ± 0.14 3.78 ± 0.35 3.1 ± 0.31 1.42 ± 0.4 2.3 ± 0.34 0.36 ± 0.09 5.06 ± 0.23 0.07 ± 0.01 1 ± 0.66 5.53 ± 0.14 2.92 ± 0.24 0.41 ± 0.11 5.23 ± 0.16 0.36 ± 0.04 0.27 ± 0.03 0.14 ± 0.04 0.2 ± 0.03 0.04 ± 0.01 0.47 ± 0.02 ND 0.1 ± 0.06 0.52 ± 0.01 0.22 ± 0.02 0.04 0.49 ± 0.02 -16.55 ± 0.63 -22.37 ± 0.18 ND -21.32 ± 0.14 -18.71 ± 0.13 ND -17.45 ± 0.93 ND ND -21.95 ± 0.14 -11.38 ± 0.17 ND 3.34 ± 0.34 3.64 ± 0.23 ND 2.05 ± 0.15 0.89 ± 0.12 ND 0.07 ± 0.01 ND ND 2.65 ± 0.11 0.33 ± 0.04 ND 0.31 ± 0.03 0.28 ± 0.02 ND 0.17 ± 0.01 0.08 ± 0.01 ND ND ND ND 0.19 ± 0.01 0.03 ND 38 816 Table 5. Slope of the relationship between sediment oxygen demand (µmol O2 m-2 h-1) and dark N2 817 efflux (µmol N2-N m-2 h-1). Regressions were forced through zero, as there would be no N2 efflux 818 when respiration equals zero. ns= not significant. Estuary All Season Slope r2 n p All 0.0170 0.743 142 <0.001 Hastings Winter Spring Summer Autumn All 0.0298 0.0241 0.0212 0.0312 0.0237 0.712 0.745 0.964 0.754 0.828 4 6 17 22 49 ns <0.05 <0.001 <0.001 <0.001 Camden Haven Winter Spring Summer Autumn All 0.0220 0.0231 0.0285 0.0148 0.0214 0.780 0.753 0.896 0.902 0.739 7 17 16 10 50 <0.05 <0.001 <0.001 <0.001 <0.001 Wallis Lake Winter Spring Summer Autumn All 0.0130 0.0114 0.0132 0.0193 0.0146 0.697 0.715 0.961 0.702 0.820 3 11 19 19 52 ns <0.01 <0.001 <0.001 <0.001 All All All All All All All All All All All All 0.0241 0.0241 0.0254 0.0254 0.0332 0.0244 0.0188 0.0211 0.0126 0.0142 0.0139 0.0165 0.697 0.801 0.790 0.440 0.536 0.741 0.934 0.661 0.717 0.889 0.712 0.711 14 12 9 27 34 7 3 12 9 9 3 3 <0.01 <0.001 <0.001 <0.05 <0.05 <0.05 ns <0.01 <0.01 <0.05 ns ns Habitat Intertidal sands (IS) Intertidal muds (IM) Subtidal sands (SS) Subtidal muds (SM) Fluvial muds and sands (FMS) Marine channel (MC) Deep subtidal muds (DM) Zostera (Z) Halophila (H) Ruppia (R) Macroalgae (MA) Posidonia (P) 819 820 821 822 Table 6. p values for test of significance between sediment oxygen demand versus dark N2 efflux slopes for individual habitats. Bold = <0.05. Abbreviations for habitats are given in Table 2. DM FMS H IM IS MA MC P R SM SS DM FMS 0.000 H 0.004 0.000 IM 0.081 0.043 0.000 IS 0.124 0.057 0.998 0.002 MA 0.220 0.659 0.000 0.017 0.024 MC 0.053 0.919 0.926 0.040 0.000 0.019 P 0.550 0.266 0.084 0.103 0.582 0.077 0.001 R 0.266 0.917 0.481 0.006 0.000 0.002 0.005 0.000 SM 0.073 0.718 0.742 0.771 0.018 0.000 0.003 0.031 0.000 SS 0.664 0.703 0.715 0.999 0.005 0.046 0.000 0.004 0.029 0.000 Z 0.246 0.327 0.389 0.220 0.205 0.145 0.054 0.003 0.000 0.032 0.000 Z 823 824 Table 7. p values for test of significance between sediment oxygen demand versus dark N2 efflux slopes for different seasons in the three estuaries Bold = <0.05. H = Hastings, CH = Camden Haven, WL = Wallis Lake. Camden Camden Camden Camden Wallis Wallis Wallis Wallis Hastings Hastings Hastings Hastings Haven Haven Haven Haven Lake Lake Lake Lake Winter Spring Summer Autumn Winter Spring Summer Autumn Winter Spring Summer Autumn H Winter H Spring 0.502 H Summer 0.114 0.478 H Autumn 0.849 0.153 0.001 CH Winter 0.350 0.691 0.800 0.036 CH Spring 0.361 0.818 0.401 0.782 0.021 CH Summer 0.322 0.321 0.409 0.085 0.050 0.001 CH Autumn 0.051 0.055 0.042 0.000 0.000 0.001 0.000 WL Winter 0.106 0.069 0.096 0.546 0.011 0.000 0.010 0.000 WL Spring 0.064 0.610 0.023 0.010 0.000 0.000 0.009 0.000 0.000 WL Summer 0.138 0.951 0.239 0.025 0.012 0.000 0.000 0.011 0.000 0.000 WL Autumn 0.152 0.274 0.319 0.449 0.165 0.001 0.001 0.035 0.069 0.002 0.002 825 42