Benthic Metabolism and Nitrogen Dynamics in an Urbanised Tidal Creek: Domination of DNRA Over Denitriﬁcation As a Nitrate Reduction Pathway

Estuarine, Coastal and Shelf Science 131 (2013) 271e281

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Estuarine, Coastal and Shelf Science

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Benthic metabolism and nitrogen dynamics in an urbanised tidal creek: Domination of DNRA over denitriﬁcation as a nitrate reduction pathway

Ryan J.K. Dunn a,*, David Robertson b,1, Peter R. Teasdale b, Nathan J. Waltham c,2, David T. Welsh b a Griffith School of Engineering, Gold Coast Campus, Griffith University, QLD 4222, Australia b Environmental Futures Centre, Gold Coast Campus, Griffith University, QLD 4222, Australia c Catchment Management Unit, Gold Coast City Council, PMB 5042 Gold Coast Mail Centre, QLD 9729, Australia article info abstract

Article history: Benthic oxygen and nutrient fluxes and nitrate reduction rates were determined seasonally under light Received 13 March 2013 and dark conditions at three sites in a micro-tidal creek within an urbanised catchment (Saltwater Creek, Accepted 26 June 2013 Australia). It was hypothesized that stormwater inputs of organic matter and inorganic nitrogen would Available online 9 July 2013 stimulate rates of benthic metabolism and nutrient recycling and preferentially stimulate dissimilatory nitrate reduction to ammonium (DNRA) over denitrification as a pathway for nitrate reduction. Storm- Keywords: waters greatly influenced water column dissolved inorganic nitrogen (DIN) and suspended solids con- benthic metabolism centrations with values following a large rainfall event being 5e20-fold greater than during the stormwater impacts nutrient fluxes preceding dry period. Seasonally, maximum and minimum water column total dissolved nitrogen (TDN) denitrification and DIN concentrations occurred in the summer (wet) and winter (dry) seasons. Creek sediments were dissimilatory nitrate reduction to highly heterotrophic throughout the year, and strong sinks for oxygen, and large sources of dissolved ammonium organic and inorganic nitrogen during both light and dark incubations, although micro-phytobenthos sub-tropical (MPB) significantly decreased oxygen consumption and N-effluxes during light incubations due to Saltwater Creek photosynthetic oxygen production and photoassimilation of nutrients. Benthic denitrification rates ranged from 3.5 to 17.7 mmol N m2 h 1, denitrification efficiencies were low (<1e15%) and denitrification was a minor process compared to DNRA, which accounted for w75% of total nitrate reduction. Overall, due to the low denitrification efficiencies and high rates of N-regeneration, Saltwater Creek sediments would tend to increase rather than reduce dissolved nutrient loads to the downstream Gold Coast Broadwater and Moreton Bay systems. This may be especially true during wet periods when increased inputs of particulate organic nitrogen (PON) and suspended solids could respectively enhance rates of N-regeneration and decrease light availability to MPB, reducing their capacity to ameliorate N- effluxes through photoassimilation. Ó 2013 Elsevier Ltd. All rights reserved.

1. Introduction Spilmont et al., 2011; Pagès et al., 2012; Dunn et al., 2012a), resulting in significant spatial and temporal variations in benthic processes The biogeochemistry of coastal waterways is influenced by many (Sundbäck et al., 2000; Welsh et al., 2000; Wilson and Brennan, 2004; competing and interacting physical and biological factors (e.g. Thornton et al., 2007; Nizzoli et al., 2007; Dunn et al., 2012a). Surface Rysgaard et al., 1995; Sundbäck et al., 2000; Bartoli et al., 2000; sediments play a significant role in the microbially mediated trans- Azzoni et al., 2001; Sakamaki et al., 2006; Dunn et al., 2009; formations of nitrogen (Rysgaard et al., 1993, 1995; Fenchel et al., 1998; Dunn et al., 2012a) and therefore understanding the processes that influence the attenuation and recycling of N-species is of * Corresponding author. Present address: Asia-Pacific ASA Pty. Ltd., P.O. Box 1679, great importance for coastal managers challenged with managing Surfers Paradise, QLD 4217, Australia. competing conservation and development land uses (Laima et al., E-mail addresses: [email protected], [email protected] (R.J.K. Dunn). 2002; Wilson and Brennan, 2004). 1 Present address: Science Museum (London), Exhibition Road, London SW7 2DD, Coastal waterways are often subject to extensive urbanisation United Kingdom. 2 Present address: Centre for Tropical Water and Aquatic Ecosystem Research (Pauchard et al., 2006; Lee et al., 2006), which typically results in (TropWATER), James Cook University, QLD 4811, Australia. major changes in both the volume and quality of stormwater runoff

(Line and White, 2007; Bratieres et al., 2008). Urban stormwater is inventory of Saltwater Creek (Tomlinson et al., 2006) in response to known to transport elevated concentrations of heavy metals public enquires regarding the health and ecology of the catchment. (Marsalek and Marsalek, 1997; Dunn et al., 2007a; Waltham et al., The study revealed considerable stress was being placed on the 2011), organic compounds (Mermillod-Blondin et al., 2005), sedi- system due to elevated sediment and nutrient loads from storm- ments (Brezonik and Stadelmann, 2002) and nutrients (Lee et al., water inputs, and concluded that future planned urban develop- 2006; Avila-Foucat et al., 2009) into receiving waters, usually un- ment in the catchment would exacerbate this pressure, and further treated and unfiltered. These anthropogenic inputs influence compromise public values and amenity of the catchment. benthic respiration rates and the nutrient status of receiving wa- The objective of this study was to present an initial assessment terways due to the increased delivery of organic matter, and dis- of the spatial and seasonal variations in benthic metabolism, solved and particulate nutrients. In some instances, increased nutrient fluxes, and rates of N-cycling processes in the intertidal nutrient loads can ultimately result in eutrophication of the sediments of Saltwater Creek and the impacts of stormwaters on receiving water body (Nixon, 1995; Taylor et al., 2005). Alterations water column nutrients, N-cycling processes and physico-chemical in concentration gradients between the water column and surface parameters. This urban catchment is representative of small coastal sediments induce change in the fluxes of oxygen and nutrients catchments throughout Moreton Bay, Australia. The study is a blue across the sediment-water interface, which can also influence rates print for similar system based nutrient process investigations, and of nitrate reduction processes within the sediment (Blackburn and aims to provide the information necessary for the future manage- Blackburn, 1993; Fenchel et al., 1998). However, the effects of urban ment of the waterway. runoff on receiving water quality are highly site specific(US EPA, 1983), making it difficult to predict the impacts or design appro- 2. Methods priate management and control programs (Brezonik and Stadelmann, 2002). 2.1. Study location Nutrient exchanges and nitrogen cycling pathways in shallow coastal systems are generally quantified by scaling up individual Saltwater Creek is a micro-tidal estuarine creek with an measurements (Eyre and Ferguson, 2005 and references therein), urbanised and modified freshwater catchment located within which are often used as inputs within system models to determine southern Moreton Bay (Australia). The creek system is approxi- allocations of management resources (Eyre and Ferguson, 2005). mately 17 km long, flowing from its headwaters in Nerang State Therefore, a good understanding of the spatial and temporal vari- Forest to the Coombabah Creek estuary confluence which connects ability of nutrient dynamics is critical if these scaled up rates are to to the Coomera River and the Gold Coast Broadwater (Fig. 1). Nat- give reliable system-wide estimates. Knowledge of spatial and ural vegetation is sparse along the estuary, and where present is temporal variations provide insight into influential factors con- dominated by mangroves (Avicennia marina, Rhizophora stylosa and trolling and maintaining benthic exchanges and nutrient cycling Aegiceras corniculatum) with saltmarsh (Sporobolus virginicus) pathways (Eyre and Ferguson, 2005). This is important for effective located downstream towards the entrance of the creek. The upper planning and integrated management tools, and allows improved estuary has several tidally connected residential canal estates predictions regarding environmental changes due to anthropo- (Benfer et al., 2007; Waltham and Connolly, 2011) and is revetted genic impacts (Sakamaki et al., 2006). with rock gabion foreshore armour for protection against erosion. Saltwater Creek is a subtropical, urbanised creek within the The Helensvale Wastewater Treatment Plant also discharged southern Moreton Bay catchment (Queensland) and the estuary is licensed treated (secondary) wastewater to the upper estuary, on part of the Moreton Bay Marine Park. In 2005, Gold Coast City the ebbing tide, until 2000 when it was decommissioned Council (local government authority) completed an environmental (Tomlinson et al., 2006). The intertidal zone of the creek is a

Fig. 1. Map (a) Australia, showing (b) southern Moreton Bay, and (c) sample sites within Saltwater Creek (: represents the location of the Helensvale Wastewater Treatment Plant discharge point). R.J.K. Dunn et al. / Estuarine, Coastal and Shelf Science 131 (2013) 271e281 273 typically turbid system (Waltham, 2002; Tomlinson et al., 2006) 2.3. Sediment and water collection and extends w10 km upstream from the creek mouth with a tidal amplitude of <1 m. Maximum water depth is 4 m and decreases in Sediment cores w15 cm in depth, for flux and process rate de- the upstream direction. The creek catchment is 30 km2 and has terminations, were hand collected at low water using plexiglass rapidly changed from a forested environment to an extensive urban (200 mm diameter 330 mm) tubes before transportation (within residential landscape, including golf course, canal estates, roads, 2 h) to the laboratory. Water used for core maintenance and incu- and commercial and industrial estates (Waltham, 2002). The creek bation was collected during the following flood tide using clean experiences a wide range of hydrodynamic conditions and sedi- (washed with deionised water [Milli-Q Element] and rinsed with ment loadings according to inflow conditions (Webster and creek water) 40 L plastic containers. Triplicate water samples were Lemckert, 2002), complicated by interactions with the Coomera collected for the determination of DIN, TDN, chlorophyll-a (chl-a) River system (Webster and Lemckert, 2002). The catchment and total suspended solids (TSS) concentrations. Surface sediment drainage network consists of overland and underground concrete light intensity at each site was measured using an LI-COR radiation pipes that deliver untreated stormwater directly to the creek. The sensor (SA: LI-192SA quantum sensor). Light intensities were local climate is characterised by warm/hot humid summers influ- measured w15 cm above the sediment surface at a fixed moni- enced by monsoonal trade winds. Thunderstorms are common toring period (i.e. w12:00 1 h). In situ water measurements were during the summer period (NovembereJanuary) often resulting in collected using a calibrated multi-probe monitoring unit (TPS 90- intense short periods of catchment freshwater flow. In contrast, FLMV, TPS). winters are typically characterised by low rainfall (average 20 mm Cores collected for the determination of sediment characteris- per month; Waltham, 2002). Daily rainfall values for the Gold Coast tics (wet-bulk density, porosity, organic matter determined as loss- region, measured at Gold Coast Seaway (Fig. 1) by the Australian on-ignition (LOI550), grain size distribution, chl-a, phaeopigment þ Bureau of Meteorology during the study year (2008) are shown in and bioavailable ammonium (NH4 bio; porewater þ exchangeable þ Fig. 2. NH4 ) concentrations were hand collected using 50 mm diameter 400 mm PVC sample tubes. Cores were immediately e e e e e e 2.2. Study design sliced into six depth horizons (0 1, 1 2, 2 4, 4 6, 6 10 and 10 15 cm) and stored in the dark (<4 C) before freezing (20 C) Three sites located approximately 2.0, 6.0 and 9.5 km upstream within 2 h of collection. of the creek mouth (Fig. 1) were sampled in early and late Autumn (March and May), winter (July) and summer (December) of 2008. 2.4. Determination of oxygen and nutrient fluxes During each survey, six sediment cores were collected at each site fl for the determination of sediment-water column uxes of oxygen, Following collection, cores were immediately returned to the S þ dissolved inorganic nitrogen (DIN; NOx ( NO2 NO3 ) and NH4 ), laboratory, carefully filled with creek water and submerged in total dissolved nitrogen (TDN) and dissolved organic nitrogen holding tubs (220 L) at seasonally measured in situ water temper- (DON), and rates of NO3 reduction processes. Cores collected for atures (23, 22, 19 and 24 1 C in autumn (March), autumn (May), sediment incubations were collected from the three sites during winter and summer, respectively). An aquarium pump and air- each seasonal event over a one week period. Water samples were stone was fitted within each core to facilitate water circulation collected, during each seasonal sediment collection period, for and aeration and triplicate cores were equilibrated under light and analysis and core incubation during the high tide period immedi- dark conditions for w12 h, at the seasonally measured in situ light ately following seasonal sediment collections. Additional water intensity (w80 mEm 2 s 1). Following equilibration, the light and samples were collected at each site, at high and low water, on the dark conditions were reversed, w40% of the water in the holding 12th May 2008 following a two week dry period and the 3rd June tanks replaced with fresh site water and the cores re-equilibrated 2008 following a large rainfall event (90 mm over previous 72 h and under the new conditions for 2 h. 65 mm in preceding 24 h period). Lastly, during the summer sample To initiate incubations, the air-stones were removed from the period, three additional sediment cores were collected at each site cores with the water lowered to below the core rims. The aquarium for the characterisation of physico-chemical surface sediment pumps were left within each core to maintain water circulation parameters. (resuspension of core sediments was avoided). Initial water samples for dissolved oxygen, DIN and TDN were then taken and the cores were then closed using floating plastic lids to prevent gaseous exchange with the atmosphere. Cores were incubated for w1.5 h and at the end of this period the aquarium pumps were stopped, the floating lids removed and final time water samples immediately collected. Flux rates (mmol m 2 h 1) were calculated from the change in water column concentrations of the individual solutes following Welsh et al. (2000).

2.5. Determination of nitrate reduction rates

Following flux incubations, the air-stones and aquarium pumps were replaced in each core, the water level in the tanks was raised to above the core tops and the cores were left to equilibrate for w 2.5 h before the determination of NO3 -reduction rates using the isotope paring technique (IPT), as modified for simultaneous determination of denitrification and dissimilatory nitrate reduction Fig. 2. Daily recorded rainfall rates for the Gold Coast Seaway 2008 (data sourced: to ammonium (DNRA). IPT core treatment and sample collection Australian Bureau of Meteorology). followed the protocol described by Dunn et al. (2012a). 274 R.J.K. Dunn et al. / Estuarine, Coastal and Shelf Science 131 (2013) 271e281

Rates of total denitrification (D14), coupled nitrification- sodium hydroxide digestion solution; 2 45 min digestions; fi fi denitri cation (DN) and denitri cation of NO3 from the overlying 120 C). Dissolved organic nitrogen (DON) was determined by dif- ¼ þ þ water (DW) were determined following Nielsen (1992). DNRA rates ference (DON TDN (NOx NH4 ). Milli-Q Element water and fi based on water column NO3 (DNRAW) were calculated from the ltered low nutrient seawater were used for all sample preparation 15 fi N-enrichment of the water column NO3 pool and the extracted and analyses. Natural ltered seawater references standards and þ fi sediment bioavailable NH4 pool (Risgaard-Petersen and Rysgaard, internal standards were used for quality assurance. Digestion ef - 1995). DNRA rates coupled to nitrification (DNRAN) were esti- ciency was established through the analysis of digested seawater mated from the rate of DNRAW and the ratio between DN and DW reference standards. Recoveries of all nutrients from the certified (Risgaard-Petersen and Rysgaard, 1995) and total rates of DNRA as standards ranged from 80 to 109%. Dissolved N2 concentrations and 29 30 15 the sum of DNRAW and DNRAN. Anammox is recognised as an the proportions of N2 and N2 and N enrichment of sediment- þ interference when using IPT that can lead to overestimates of NH4 bio pools were analysed at the National Environmental denitrification rates, as it also generates labelled N2 species Research Institute (Silkeborg), Denmark. 15 following NO3 additions (Risgaard-Petersen et al., 2003). How- ever, in shallow water sediments anammox is a minor source of N2 2.7. Macrofauna dynamics compared to denitrification (Dalsgaard et al., 2005; Burgin and Hamilton, 2007), especially in tropical systems (Dong et al., 2011). Following determinations of nitrate reduction rates, sediments Therefore, the authors believe that reported estimates of denitri- from each incubation core plus the corresponding KCl-sediment fication are valid, although it should be noted that the term deni- slurry were pooled and sieved (250 mm mesh) to recover burrow- trification here also includes a small portion of N2 produced via ing macrofauna. Macrofauna were rinsed with freshwater and anammox. preserved in 70% ethanol, counted and identified.

2.6. Sample handling and analytical techniques 2.8. Statistical analyses

fl Sediment wet-bulk density, porosity and LOI550 were deter- Variation in ux and nitrate reduction rates were analysed using mined following Dunn et al., 2007b. The proportion of clay and silt three-way analysis of variance (ANOVA) with light/dark conditions, (<63 mm), sand (180 < x > 63 mm) and gravel (>180 mm) sediment month (season) and site fixed, and interaction included. Season was fractions were determined by dry sieving and are expressed as a a fixed factor here as each survey represents a single point in time, percent dry weight. Sediment chl-a and phaeopigment concen- with the timing of each survey representative of known maximal trations were determined following freeze drying and acetone differences in local environmental conditions (i.e. temperature, extraction of 1 cm3 aliquots of surface sediment following Lorenzen rainfall, flow). Prior to analyses the assumption of homogeneity of þ variances was tested using box and whisker plots, before and after (1967). Sediment NH4 bio concentrations were determined following extraction of 1 cm3 of homogenised sediment for 24 h in transformation (Quinn and Keough, 2008). Variances were best 9 ml 2 M KCl (Nizzoli et al., 2005). stabilised with a log (x) transformation. Post hoc comparisons were Water samples were collected using acid washed (10% v/v HCl), performed using Tukey’s HSD. Correlations between physico- sample rinsed, low density polyethylene sample bottles (Nalgene). chemical sedimentary conditions, faunal communities, solute Dissolved inorganic and total nutrient samples were immediately fluxes and nitrate reduction rates were analysed using Pearson filtered through pre-washed, pre-ashed 25 mm GF/F filters and correlation analysis (2-tailed). Statistical significance was deter- stored frozen (20 C) until analysed. Chl-a samples were filtered mined at a ¼ 0.05. All statistical analyses were performed using through 25 mm pre-washed, pre-ashed GF/C filters immediately SPSS Windows (SPSS Inc., version 19). following water collection and the filters stored frozen until analysed (APHA, 2005). TSS samples were filtered through pre-washed, 3. Results pre-ashed 47 mm GF/F filters. During sediment incubation, water samples were collected using acid washed and Milli-Q element 3.1. Sediment characteristics water rinsed 60 ml plastic syringes and tubing. Nutrient samples were filtered (GF/F) and stored frozen (20 C) until analysis. Dis- Sediment characteristics were relatively homogenous over the solved oxygen samples were carefully transferred to gas-tight 12 ml 0e15 cm depth horizon sampled and are therefore expressed as glass vials (Exetainer, Labco Ltd.), fixed with 100 ml of manganous depth integrated means, with the exception of Chl-a and phaeo- sulfate and alkali-iodide-azide solution (APHA, 2005), stored at 4 C pigment concentrations, which were only determined in the sur- and analysed within 48 h according to the Winkler titration (Azide- face 1 cm (Table 1). Sediments at all sampling sites were dominated modification method, APHA, 2005). by the <63 mm and 63e180 mm size fractions, which together DIN concentrations were directly determined by an automated accounted for 83e90% of the total particle size distribution. Sedi- nutrient analyser (Easychem Plus Random Access analyzer, Systea ment organic matter content (LOI550), increased from 5.8 1.9% dry Analytical Technologies). TDN concentrations were determined as wt at site 1 (closest to creek mouth) to 8.0 2.2 at site 3 (most þ þ NOx (NO3 NO2 ) following digestion (potassium persulfate/ upstream estuary site). Sediment NH4 bio concentrations ranged

Table 1 Mean (SD) depth integrated (0e15 cm) sediment characteristics and surface (0e1 cm) concentrations of Chl-a and phaeopigment concentrations in the estuary. n ¼ 18 for þ ¼ particle size distributions, density, porosity, LOI550 and NH4 bio data and n 3 for Chl-a and phaeopigment content. þ Site Particle size distribution (%) Density Porosity LOI550 NH4 bio Chl-a Phaeopigment <63 mm63e180 mm >180 mm(gcm3) (nmol g dry wt 1) (mg g dry wt 1) (mg g dry wt 1)

130 9.4 60.2 5.1 11.9 5.4 1.53 0.13 69.0 4.4 8.0 0.4 238 145 1.5 0.4 2.1 0.5 2 56.4 7.5 31.9 7.2 11.5 3.4 1.80 0.18 61.3 7.1 6.2 1.6 201 73 0.6 0.1 0.8 0.1 3 43.0 12.9 40.1 9.8 16.6 7.6 1.55 0.17 52.7 6.4 5.8 1.9 198 102 0.9 0.8 1.3 1.1 R.J.K. Dunn et al. / Estuarine, Coastal and Shelf Science 131 (2013) 271e281 275

from 198 102 to 238 145 nmol g dry wt 1 at site 3 and 1, Table 3 respectively. Chl-a and phaeopigment concentrations were greatest Water column nutrient concentrations and physico-chemical parameters at each sampling site at high and low tide during dry conditions (12th May 2008; no rainfall at the seaward site 1. in preceding 2 weeks) and wet conditions following a significant rainfall event (3rd June 2008; 90 mm over 72 h, 65 mm in preceding 24 h). PN represents particulate 3.2. Temporal variations in water column characteristics nitrogen, all other abbreviations as in Table 3. þ Sample site, NOx NH4 DIN DON TDN PN TSS Seasonal water column DIN concentrations ranged between conditions (mM) (mM) (mM) (mM) (mM) (mM) (mg L 1) 1.54 0.68 and 10.33 0.88 mM(Table 2). Maximum DIN con- and tide centrations were measured during the summer wet season, coin- Dry ciding with reduced salinity. DIN concentrations were typically Site 1 high 0.6 0.5 1.1 28.2 29.1 31.3 19.5 Low 0.6 0.4 0.9 28.2 29.1 31.3 23.5 dominated by the contribution of NOx, which represented 63.1% of the total DIN pool. However, DON was the dominant dissolved ni- Site 2 high 0.6 0.8 1.4 25.4 26.8 28.8 30.5 e Low 0.5 0.8 1.3 29.7 32.9 35.4 29.4 trogen species, representing 89.4 98.2% of TDN, with the lowest Site 3 high 0.6 2.2 3.1 30.8 33.7 35.4 25.5 contribution of DON to the TDN pool recorded during the summer Low 1.0 2.6 3.8 29.1 32.9 36.3 28.5 survey, despite the highest measured DON concentration also Wet occurring in summer. Site 1 high 18.1 5.4 25.7 15.8 41.5 44.7 97.0 Dry and wet event sampling showed that following a significant Low 19.9 5.3 24.1 17.2 41.4 44.5 83.0 rainfall event in June (90 mm over 72 h, 65 mm in preceding 24 h), Site 2 high 31.9 3.3 36.0 19.1 55.1 59.2 263.0 þ Low 34.1 3.2 38.1 4.0 42.1 50.6 250.0 concentrations of DIN, NH and NOx were 5e25-fold greater than 4 Site 3 high 33.4 6.3 40.3 6.1 46.4 58.4 303.3 those in May or July, with the highest concentrations measured at Low 27.3 7.4 35.5 11.6 47.1 57.5 283.3 the upstream sites (Table 2 and 3). In contrast, DON concentrations during this wet event were 50% or less of those measured during the May or July surveys, with the lowest DON concentrations between light and dark oxygen fluxes ranged between 400 (site 3, measured at sites 2 and 3. Similarly, concentrations of total sus- autumn; May) to 2000 mmol m 2 h 1 (site 1, autumn; May). Rates of pended solids and particulate nitrogen were also elevated by gross production were highest at site 1 and lowest at site 3, except approximately 2 and 10-fold respectively, under wet compared to during summer, where the highest rate occurred at site 3. dry conditions, with this effect again most evident at the upstream þ Sediments at all sites in all seasons were a net source of NH4 sampling sites (Table 3). þ fl during dark incubations and sediment dark NH4 ef uxes (Fig. 5) and oxygen demands were significantly correlated (r ¼ 0.682, 3.3. Burrowing macrofauna dynamics p < 0.001). During light incubations, sediments were either small þ fl sinks for NH4 or showed much reduced rates of ef ux. Seasonal þ fl fi In total 420 individuals were collected, consisting of three spe- NH4 uxes showed signi cant seasonal trends with highest ef- cies of surface and sub-surface deposit and filter feeding macro- fluxes measured during summer and also significant differences faunal groups: worm (Simplisetia aequisetis), amphipod between sampling sites with the sediments of sites 2 and 3 being þ þ (Victoriopisa australianesis), and crab (Heloecius cordiformis). Mac- greater sources of NH4 compared to site 1. Comparable to NH4 rofauna densities ranged from 0 to 605, 0 to 446 and 0 to 64 in- fluxes, sediments were also sources of NOx under dark conditions 2 2 dividuals m (ind m ) for worms, amphipods and crabs, and small sinks or greatly reduced sources of NOx under light respectively. The highest combined abundance occurred in summer conditions (Fig. 5). NOx fluxes also demonstrated significant sea- (Fig. 3). V. australiensis dominated the surface sediments at sites 1 sonal differences (Table 4) with highest effluxes measured in and 2 accounting for 45 and 50% of the total macrofaunal density, summer and the lowest in the autumn and winter periods. As both þ respectively. In comparison, at site 3 the abundance of S. aequisetis NH4 and NOx showed similar spatial and temporal trends, overall accounted for 56% of the total abundance. DIN fluxes followed the same pattern with sediments being large sources of DIN under dark conditions, slight sinks or greatly 3.4. Sediment-water column oxygen and dissolved inorganic reduced sources of DIN under light conditions with the highest DIN nutrient fluxes effluxes occurring in summer and the lowest in winter (data not shown). However, fluxes of DIN were small compared to DON, Sediments at all sites were consistent sinks for water column which represented 50e69% of the annual TDN efflux. Sediments dissolved oxygen during both light and dark incubations (Fig. 4), were significant sources of DON in all seasons, at all sites during although rates of oxygen consumption were significantly lower both light and dark incubations. Maximum DON effluxes occurred during light compared to dark incubations (Table 4). Dark oxygen in winter (July) and lowest in autumn (March), although the fluxes varied from 2900 to 2100, 2200 to 1900 and 4200 maximum mean contribution of DON to dark TDN effluxes (87%) to 1700 mmol m 2 h 1 at sites 1, 2 and 3, respectively. Estimates of occurred in May, with the minimum contribution (42%) measured gross community primary production calculated by difference in summer (December). As a result of the constant efflux of DON,

Table 2 Seasonal variations in water column dissolved nitrogen concentrations and physico-chemical parameters. All data are mean values (SD) (n ¼ 3). Abbreviations: Aut ¼ Autumn, Win ¼ winter, Sum ¼ summer, DIN ¼ dissolved inorganic nitrogen, DON ¼ dissolved organic nitrogen, TDN ¼ total dissolved nitrogen, chl-a ¼ chlorophyll-a, Temp. ¼ temperature and TSS ¼ total suspended sediments.

Season Dissolved nutrients Physico-chemical parameters m þ m m m m m 1 & 1 NOx ( M) NH4 ( M) DIN ( M) DON ( M) TDN ( M) Chl-a ( gL ) pH Temp. (oC) Salinity ( ) TSS (mg L ) Aut (Mar) 1.87 0.13 1.67 0.50 3.54 0.37 44.0 3.2 47.2 3.9 3.10 0.31 8.1 0.1 24.1 0.2 28.4 0.3 28.3 2.1 Aut (May) 0.81 0.44 0.74 0.45 1.54 068 55.2 4.4 56.2 3.9 2.78 1.10 8.2 0.1 22.3 0.1 27.8 0.2 30.8 4.2 Win 4.68 0.30 0.57 0.34 5.25 0.31 44.8 5.6 48.6 4.2 2.92 0.49 8.0 0.1 20.4 0.3 28.2 0.3 26.3 2.7 Sum 7.23 0.31 5.25 0.31 10.33 0.88 69.1 9.5 77.3 7.4 4.01 0.74 7.8 0.2 26.3 0.1 25.2 0.2 33.6 5.1 276 R.J.K. Dunn et al. / Estuarine, Coastal and Shelf Science 131 (2013) 271e281

Fig. 4. Seasonal sediment-water column oxygen fluxes during a) light and dark conditions (negative values indicate oxygen consumption and flux of oxygen into the sediment) and b) gross primary production at each sample site based on mean light and dark oxygen fluxes (n ¼ 3).

Fig. 3. Mean (SD) a) seasonal abundance, and b) annual site abundance of macro- ¼ fauna species across all sample cores (n 72). Species are Heloecius cordiformis (crab), fi Victoriopisa australianesis (amphipod) and Simplisetia aequisetis (worm). Nitri cation was the principal source of nitrate fuelling nitrate reduction processes in the sediment (Figs. 6 and 7), with DN and DNRAN accounting for 81.2 14.6% and 74.4 20.9% of average sediments at all sites were net sources of TDN to the water column annual nitrate reduction under light and dark conditions, respec- in all seasons during both light and dark incubations. TDN effluxes tively. In contrast, rates of DW and DNRAW were of lower impor- showed significant (p < 0.01) differences between light and dark tance, and water column nitrate was only a significant source of conditions with larger TDN effluxes occurring under dark condi- nitrate for denitrification and DNRA within the creek sediments in tions (Table 4 and Fig. 5), although this was largely due to the DIN summer (Figs. 6 and 7), which corresponded with higher NOx component. concentrations in the water column (Table 2). Sediment denitrification efficiencies calculated as N2eN/(N2eN þ DIN efflux) 100 3.5. Nitrate reduction processes (Eyre and Ferguson, 2002) ranged between 0.72 and 79.1% with an overall combined mean of 14.9 17.3% across all sampling sites and fi Total nitrate reduction rates varied between 9.2 mmol-N m 2 h 1 seasons. Lower ef ciencies generally corresponded with higher fi (site 1 autumn, dark incubation) to 110.6 mmol-N m 2 h 1 (site 2 sediment oxygen demands and greater ef ciencies were observed summer, light incubation) (data not shown) with a significant with lower sediment oxygen demands. Fig. 8 illustrates mean fi fi seasonal trend apparent with the highest rates measured in sum- denitri cation ef ciencies for individual incubation cores at sediment oxygen demands of <1000, 2000 < x > 1000 and >2000 mmol mer (Table 4). DNRA was the dominant pathway for nitrate 2 1 reduction at all sampling sites in all seasons (Figs. 6 and 7), and O2 m h , respectively. accounted for 64.9 to 81.8 and 70.1 and 86.0%, respectively, of light and dark total nitrate reduction rates. The maximum contribution 4. Discussion of DNRA to total nitrate reduction was measured during the summer survey when DNRA accounted for on average 80.3 and 86.0% of 4.1. Water column nutrient concentrations total light and dark nitrate reduction, respectively, across the sample sites. Total denitrification rates ranged between The seasonal pattern for water column DIN concentrations was 3.5 0.5 mmol-N m 2 h 1 to 17.7 5.5 mmol-N m 2 h 1 (Fig. 6), comparable with those recorded in previous studies within with significant (p < 0.001) differences measured between light Southern Moreton Bay (Tomlinson et al., 2006; Dunn et al., 2007c, and dark conditions. Mean total DNRA ranged between 2012a; Eyre et al., 2010a, 2012b). Elevated concentrations coincided 5.1 6.2 mmol-N m 2 h 1 to 92.3 94.6 mmol-N m 2 h 1 (Fig. 7), with summer wet season conditions characterised by greater with an apparent significant seasonal trend with highest rates external inputs of stormwater. DON was the dominant component recorded in summer (Table 4). of the water column TDN pool and also dominated sediment-water R.J.K. Dunn et al. / Estuarine, Coastal and Shelf Science 131 (2013) 271e281 277

Table 4 Summary of results of three-way ANOVA for seasonal solute flux and nitrate reduction rate measurements among sample sites under light and dark conditions. Significant outcomes (p < 0.05) are highlighted in bold. Abbreviations: DF ¼ degrees of freedom, TNR ¼ total nitrate reduction (denitrification þ DNRA), D14 ¼ total denitrification, DW ¼ denitrification based on nitrate diffusing from the overlying water column, DN ¼ denitrification coupled to nitrate production via nitrification in the sediment, DNRAT ¼ total DNRA, DNRAW ¼ DNRA based on nitrate diffusing from the overlying water column, DNRAN ¼ DNRA coupled to nitrate production via nitrification in the sediment; all other abbreviations as in legend to Table 3.

DF Sediment-water column ﬂuxes Nitrate reduction rates þ O2 NH4 NOx DIN DON TDN TNR D14 DW DN DNRAT DNRAW DNRAN Light 1 <0.001 0.063 0.186 <0.001 0.247 0.006 0.548 <0.001 0.391 0.003 0.773 0.646 0.549 Month 3 0.023 0.004 <0.001 <0.001 0.064 0.038 0.031 0.051 <0.001 0.518 <0.001 <0.001 0.028 Site 2 0.133 0.007 0.537 0.059 0.061 0.181 0.212 0.087 0.352 0.084 0.245 0.023 0.249 L M 3 0.280 0.379 0.279 0.090 0.951 0.258 0.345 0.284 0.895 0.379 0.291 0.236 0.312 L S20.008 0.290 0.998 0.840 0.684 0.876 0.449 0.213 0.755 0.219 0.088 0.135 0.090 M S 6 0.237 0.780 0.409 0.301 0.418 0.317 0.141 <0.001 0.020 0.002 0.297 0.845 0.315 L M S 6 0.733 0.343 0.903 0.563 0.071 0.143 0.841 0.417 0.964 0.394 0.524 0.904 0.420

column nutrient effluxes to a similar extent, indicating that under can be driven by the volume of stormwater flow emanating from dry conditions the composition of the water column TDN pool was the catchment. Additionally, given that in Southern Moreton Bay, driven by nutrient cycling processes in the sediment. w30% of the annual regional rainfall falls between December and In contrast, following a significant rainfall event in June 2008, February (summer) and rainfall events much larger than the DIN concentrations in the creek waters increased dramatically by studied June event are also common around this period (e.g. events up to 25-fold, to values that are the highest recorded concentra- in late November and early January of the study year; Fig. 2), the tions for Southern Moreton Bay (Tomlinson et al., 2006; Dunn et al., potential influence of stormwaters on physico-chemical conditions 2007c, 2012a; Eyre et al., 2010a, 2012b). This increase in DIN con- in Saltwater Creek could conceivably be even greater than those centrations was accompanied by substantial increases in TSS and measured here, which poses greater challengers for managers. particulate nitrogen concentrations, and reduced salinity and DON Nutrient inputs associated with stormwaters following heavy concentrations. These dramatic changes, clearly demonstrate that rainfall have been linked to increased phytoplankton and algal during this wet period, stormwater run-off rich in DIN, suspended growth in the nearby Coomera and Pimpama River estuaries (Eyre solids and particulate nitrogen was a major influence on overall et al., 2010b). However, given the relatively high TSS concentrations physico-chemical conditions in the creek waters. Similar patterns which persisted throughout the year in Saltwater Creek and the have been observed elsewhere (e.g. Goonetilleke et al., 2005; Flint very high TSS concentrations associated with stormwater inputs, it and Davis, 2007; Beck and Birch, 2012; Sood et al., 2012) supporting is likely that primary production in the water column is light a model that the physico-chemical conditions in the water column limited and that benthic production is essentially limited to shallow

fl þ þ Fig. 5. Seasonal sediment-water column uxes of nitrate nitrite (NOx), ammonium (NH4 ), dissolved organic nitrogen (DON) and total dissolved nitrogen (TDN; þ þ þ fl NOx NH4 DON) at each sample site during light and dark conditions. Negative values indicate consumption of the solute by the sediment (in ux) and positive values production of the solute by the sediment (efflux). Data are mean values and error bars indicate the standard deviation of the mean flux (n ¼ 3). 278 R.J.K. Dunn et al. / Estuarine, Coastal and Shelf Science 131 (2013) 271e281

fi Fig. 6. Seasonal rates of denitri cation based on nitrate diffusing from the water Fig. 7. Seasonal rates of DNRA based on nitrate diffusing from the water column column (DW), denitrification coupled to nitrification (DN), and total denitrification rates (DNRAW), DNRA coupled to nitrification (DNRAN), and total DNRA (DNRA14) at each (D14) at each sample site during light and dark conditions. Data are mean values and sample site during light and dark conditions. Data are mean values and error bars ¼ error bars indicate the standard deviation of the mean rate (n 3). indicate the standard deviation of the mean rate (n ¼ 3). subtidal and intertidal zones during periods of low water. Conse- quently, the bulk of the DIN and suspended solids entering the creek in the stormwaters are likely to be exported to the downstream Gold Coast Broadwater, where TSS may negatively impact on seagrass communities (McClennan and Sumpton, 2005; Pagès et al., 2012).

4.2. Benthic metabolism and nutrient ﬂuxes

The sediments throughout the creek system were a sink for water column dissolved oxygen under both light and dark conditions, as has also been reported in intertidal areas of other east coast Australian estuaries (Eyre and Ferguson, 2002, 2005; Qu et al., 2006; Dunn et al., 2012a). These oxygen dynamics are typical of sediments receiving high organic matter loads from external sources (organic matter inputs > local production ¼ net heterotrophic conditions (Viaroli et al., 2004)) and may at least in part be due to organic matter loading from urban stormwater. This sustained sediment oxygen demand during both light and dark conditions may negatively impact the water column, in particular, contributing to the low dissolved oxygen saturation values Fig. 8. Mean denitrification efficiencies for individual incubation cores at categorised measured in this creek system (Waltham, 2002; Tomlinson et al., 2 1 sediment oxygen demands of <1000 mmol O2 m h (n ¼ 6), 2000 < x > 1000 mmol O2 2 1 2 1 2006). Sustained sediment oxygen demand and reduced water m h (n ¼ 29) and >2000 mmol O2 m h (n ¼ 13). Error bars indicate the standard column oxygen concentrations would also favour anaerobic deviation of the mean denitrification efficiencies. R.J.K. Dunn et al. / Estuarine, Coastal and Shelf Science 131 (2013) 271e281 279 microbial metabolisms and the accumulation of the toxic reduced the % contribution of anamox to N2 production was typically below metabolites (e.g. sulfides and Fe(II)) within the sediment. The dis- 1% (Nicholls and Trimmer, 2009). Therefore, we are confident that tribution of these compounds, especially sulfides, is a major our denitrification measurements provide a true reflection of N- determinant of benthic fauna distributions and could be respon- loss as N2 from the Saltwater creek sediments. sible for the low abundance, biomass and biodiversity of infauna The creek sediments also had low denitrification efficiencies, observed in the creek sediments (Cosser, 1989). which were again comparable with estuaries elsewhere (e.g. Micro-phytobenthos were present at the sediment surface, as Berelson et al., 1998; Heggie et al., 1999; Eyre and Ferguson, 2002; indicated by chl-a concentrations and shifts in sediment oxygen Cook et al., 2004; Dunn et al., 2012a), presumably as a result of high fluxes measured during light compared to dark conditions, carbon loadings. Large decreases in denitrification efficiency have although oxygen fluxes remained negative during light conditions. been reported at carbon decomposition rates >w1500 mmol CO2 2 1 2 1 The consistent consumption of oxygen by the sediment during light m h (Eyre and Ferguson, 2002) and w1250 mmol CO2 m h incubations reflects the high inputs of organic matter, which drive (Berelson et al., 1998). Therefore, the high carbon loadings within respiratory processes and the high turbidity of the water column. the surface sediments of Saltwater Creek, as indicated by elevated Increased turbidity limits light availability for photosynthesis and it LOI550 values and oxygen demands, presumably contributed to the is likely that in these intertidal sediments microalgal production is observed low denitrification efficiencies. During this study the essentially limited to the low water period, during sediment lowest benthic denitrification efficiencies measured in individual emersion. However, this production would not contribute to re- incubation cores coincided with those having the highest sediment oxygenation of the water column. It should be noted that data oxygen demands and denitrification efficiency decreased steeply in from our light incubated cores represent a best case scenario, as the cores with oxygen demands >2000 mmol m 2 h 1 (Fig. 8). This bulk of the sub-tidal creek sediments are unlikely to receive suffi- hypothesis is supported by the significant increase in rates of cient light to support microalgal communities and would therefore coupled nitrification-denitrification that occurred during light in- act as strong constant sinks for oxygen regardless of prevailing light cubations, which indicates that nitrification and hence denitrifi- conditions at the water surface. cation rates in the Saltwater Creek sediments were limited by þ Heterotrophic metabolism in the sediments drove sustained oxygen rather than NH4 availability and that this limitation can be effluxes of dissolved nitrogen species during all seasons at all sites, offset by photosynthetic oxygen production, as has been observed although photoassimilation of N did significantly reduce effluxes in other strongly heterotrophic sediments colonised by benthic during light incubations. Highest DIN and TDN effluxes occurred microalgae (An and Joye, 2001; Dunn et al., 2012a). However, the during summer and effluxes showed similar magnitudes and sea- dominance of DNRA over denitrification as a sink for nitrate would sonal dynamics to those reported for other turbid east coast also substantially contribute to the low denitrification efficiencies Australian estuaries (Qu et al., 2005; Eyre et al., 2010a; Dunn et al., in the sediments through competition between denitrification and 2012a). DON was the dominant contributor to sediment TDN ef- DNRA for NOx, although DNRA itself is also favoured by high fluxes, accounting for on average 50e70% of the TDN efflux across sediment carbon loadings and metabolic rates (Tiedje, 1988; sites. Model and mesocosm simulations (Blackburn and Blackburn, Christensen et al., 2000; Nizzoli et al., 2006). Consequently, re- 1993; Sloth et al., 1995) indicate that high contributions of DON to ductions in carbon loadings to the creek from stormwater would overall TDN effluxes are characteristic of sediments that receive presumably enhance the sediment denitrification efficiency and elevated organic matter loads that are mineralised primarily at/or the natural removal of nitrogen within the creek sediments, both by near to the sediment-water interface and not mixed into the deeper reducing sediment respiration rates and by favouring denitrifica- sediment layers by bioturbation. Thus, the high TDN effluxes and tion over DNRA as a pathway for nitrate reduction. high contribution of DON to these effluxes may reflect the high DNRA was the dominant pathway for nitrate reduction in all organic loads the sediments receive from stormwater inputs and seasons at all sites, accounting for w75% of total nitrate reduction. the limited abundance and biodiversity of macrofauna, which Rates of total DNRA, DNRAW and DNRAN all showed the same sea- limits mixing of this organic matter into the deeper sediment sonal pattern with highest rates occurring during summer. These strata. Consequently, organic matter is mineralised at or close to the results are in agreement with those other recent studies of tropical sediment-water interface resulting in a high proportion of the early and sub-tropical estuaries (Dong et al., 2011; Dunn et al., 2012a; þ mineralization products such as DON and NH4 escaping to the Molnar et al., 2013) and support the hypothesis that DNRA is a water column by diffusion before they can be further metabolized very much more important process in the N-dynamics of tropical (Blackburn and Blackburn, 1993; Sloth et al., 1995). and sub-tropical compared to temperate estuarine sediments. Several factors have been proposed to favour DNRA over denitrifi- 4.3. Nitrate reduction processes cation as a pathway for nitrate reduction including; high temperature, high ratios of labile organic carbon to NO3 ratios, elevated rates fi Denitri cation rates in the Saltwater Creek sediments were of benthic metabolism, low NO3 availability, and reduced, especially comparable to those measured in western Moreton Bay (w6e sulphidic sediment conditions (see Nizzoli et al., 2006 and refer- w27 mmol-N m 2 h 1; Ferguson et al., 2007), Coombabah Lake (<1 ences therein), although the influence of the factors is difficult to to 6 mmol-N m 2 h 1; Dunn et al., 2012a) and the (regional scale) separate as they are often coincident (Nizzoli et al., 2006). Our data Brunswick estuary (w1e58 mmol-N m 2 h 1; Eyre and Ferguson, are in general agreement with these regulatory mechanisms as 2005). Anammox was not directly measured, but would be ex- Saltwater Creek water temperatures were typically above 20 C and pected to make only a minor contribution to overall N2 production benthic respiration rates were elevated throughout the year, and the in the studied estuary. For example, in a study of 40 sites in 9 es- largest dominance of DNRA over denitrification occurred during tuaries in southern England, Nicholls and Trimmer (2009) found summer when temperature and rates of benthic respiration were that in sediment slurries amended with 15N substrates, that ana- maximal and the sediment redox conditions would likely be most mox accounted for only 1e11% of total N2 production. In this study, reduced. the % contribution of anammox to N2 production increased with Nitrification was the primary source of nitrate fuelling nitrate water column nitrate concentration over the range of 3e790 mM. reduction in the Saltwater Creek sediments with DN and DNRAN However, at the water column nitrate concentrations in the 1e7 mM accounting for w75% of total nitrate reduction under both light and range measured in Saltwater Creek during our seasonal samplings, dark conditions. As discussed earlier, rates of DN were significantly 280 R.J.K. Dunn et al. / Estuarine, Coastal and Shelf Science 131 (2013) 271e281 stimulated during light incubations indicating that microalgal Griffith Centre for Coastal Management (Griffith University, Gold photosynthesis played a regulatory role and oxygen rather than Coast Campus). þ NH4 availability was limiting nitrification rates. Thus, despite their relatively low biomass benthic microalgae still played a major role References in regulating N-fluxes and N-cycling processes through the photoassimilation of N-species and the production of oxygen. American Public Health Association (APHA), American Water Works Association fi and Water Pollution Control Federation, 2005. Standard Methods for Exami- Although nitri cation was the major source of nitrate for deni- nation of Water and Wastewater, twenty-first ed. American Public Health trification, rates of denitrification dependent on nitrate diffusing Association, Washington, D.C., p. 1368. fi fi from the overlying water column (DW) showed significant seasonal An, S., Joye, S.B., 2001. Enhancement of coupled nitri cation-denitri cation by variations, with maximum rates measured during summer coin- benthic photosynthesis in shallow estuarine sediments. Limnology and Oceanography 46, 62e74. ciding with the maximum measured concentrations of NOx in the Avila-Foucat, V.S., Perrings, C., Raffaelli, D., 2009. An ecologicaleeconomic model for water column. Such regulation of DW by water column nitrate catchment management: the case of Tonameca, Oaxaca, México. Ecological e concentration would be expected and has frequently been observed Economics 68, 2224 2231. Azzoni, R., Giordani, G., Bartoli, M., Welsh, D., Viaroli, P., 2001. Iron, sulphur and in marine sediments (e.g. Rysgaard et al., 1995; Christensen et al., phosphorus cycling in the rhizosphere sediments of a eutrophic Ruppia cirrhosa 2000; Nizzoli et al., 2006; Hietanen and Kuparinen, 2008), as the meadow (Valle Smarlacca, Italy). Journal of Sea Research 45, 15e26. fl rate of supply of nitrate from the water column to the sediment Bartoli, M., Nizzoli, D., Welsh, D.T., Viaroli, P., 2000. Short-term in uence of fi recolonisation by the polydhaete worm Nereis succinea on oxygen and nitrogen denitri cation zone largely depends on the concentration differ- fluxes and denitrification: a microcosm simulation. Hydrobiologia 431, 165e174. ence between the two zones. In Saltwater Creek this effect would Beck, H.J., Birch, G.F., 2012. Metals, nutrients and total suspended solids discharged during different flow conditions in highly urbanised catchments. Environ- be amplified, as peak water column NOx concentrations occurred in mental Monitoring and Assessment 184, 637e653. summer when sediment oxygen demand was maximal and there- Benfer, N.P., King, B.A., Lemckert, C.J., 2007. Salinity observations in a subtropical fore oxygen penetration into the sediment minimal (Fenchel et al., estuarine system on the Gold Coast, Australia. Journal of Coastal Research 50, 1998). Such conditions would further enhance the concentration 646e651. Berelson, W.M., Heggie, D., Longmore, A., Kilgore, T., Nicholson, G., Skyring, G., 1998. gradient of nitrate and therefore the diffusion rate of nitrate from Benthic nutrient recycling in Port Phillip bay, Australia. Estuarine Coastal and the water column to the sediment denitrification zone. It would Shelf Science 46, 917e934. also be expected that, during wet periods, increased inputs of ni- Blackburn, T.H., Blackburn, N.D., 1993. Rates of microbial processes in sediments. e trate to the creek during stormwater flow would greatly stimulate Philosophical Transactions of the Royal Society London A 344, 49 58. Bratieres, K., Fletcher, T.D., Deletic, A., Zinger, Y., 2008. Nutrient and sediment rates of DW, as following the sampled high rainfall event in June the removal by stormwater biofilters: a large-scale design optimisation study. e water column NOx concentrations were 4e34-fold higher than Water Research 42, 3930 3940. those measured during the seasonal sampling campaigns. How- Brezonik, P.L., Stadelmann, T.H., 2002. Analysis and predictive models of stormwater runoff volumes, loads, and pollutant concentrations from watersheds in the ever, due to the dominance of DNRA as a pathway for nitrate Twin Cities metropolitan area, Minnesota, USA. Water Research 36, 1743e1757. fi reduction in the creek sediments, this increase in DW would have Burgin, A.J., Hamilton, S.K., 2007. Have we overemphasized the role of denitri ca- little impact on the eutrophication status of the creek or the tion in aquatic ecosystems? A review of nitrate removal pathways. Frontiers in Ecology and the Environment 5, 89e96. nutrient loads transported to downstream water bodies. Christensen, P.B., Rysgaard, S., Sloth, N.P., Dalsgaard, T., Schwaerter, S., 2000. Sedi- In terms of eutrophication, DN can be considered as the sediments ment remineralization, nutrient fluxes, denitrification and dissimilatory nitrate reduction in an estuarine fjord with sea cage trout farms. Aquatic Microbial capacity to eliminate internal nitrogen loads and DW as the sedi- Ecology 21, 73e84. ments capacity to eliminate external nitrate loads during transport Cook, P.L.M., Eyre, B.D., Leeming, R., Butler, E.C.V., 2004. Benthic fluxes of nitrogen in (Nizzoli et al., 2006). Whereas, DNRAW represents an input of N to the tidal reaches of a turbid, high-nitrate sub-tropical river. Estuarine Coastal the sediment compartment, as nitrate diffusing into the sediment is and Shelf Science 59, 675e685. Cosser, P.R., 1989. Water quality, sediments and the macroinvertebrate community converted to ammonium and DNRAN constitutes a futile cycle, as of residential canal estates in south-east Queensland, Australia: A multivariate nitrate produced in the sediment from ammonium by nitrification is analysis. Water Research 23, 1087e1097. simply recycled back to ammonium in the sediment by DNRA Dalsgaard, T., Thamdrup, B., Canfield, D.E., 2005. 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Indeed, as benthic rates of total denitrifica- the time-integrated, in situ DGT technique by monitoring changes in heavy tion were small compared to the benthic effluxes of DIN and TDN, the metal concentrations in estuarine waters in response to natural and anthro- e net effect of N-cycling in the creek sediments would be to enrich the pogenic processes. Environmental Pollution 148, 213 220. Dunn, R.J.K., Lemckert, C.J., Teasdale, P.R., Welsh, D.T., 2007b. Distribution of nu- overlying waters with dissolved N-species during transport. This trients in surface and sub-surface sediments of Coombabah Lake, southern may be especially true during wet periods for two principal reasons. Moreton Bay (Australia). Marine Pollution Bulletin 54, 606e614. Firstly, as stormwater increased particulate N concentrations in the Dunn, R.J.K., Ali, A., Lemckert, C.J., Teasdale, P.R., Welsh, D.T., 2007c. Short-term variability of physico-chemical parameters and the estimated transport of water column approximately 2-fold, sedimentation of these partic- filterable nutrients and chlorophyll-a in the urbanised Coombabah Lake and ulates would be expected to further stimulate benthic metabolism Coombabah Creek system, southern Moreton Bay, Australia. Journal of Coastal and nutrient effluxes. Secondly, stormwater increased TSS concen- Research 50, 1062e1068. e Dunn, R.J.K., Welsh, D.T., Jordan, M.A., Teasdale, P.R., Lemckert, C.J., 2009. Influence trations in the creek waters 3 10-fold, which would further decrease of natural amphipod (Victoriopisa australiensis) (Chilton, 1923) population light availability to benthic microalgae and therefore their capacity to densities on benthic metabolism, nutrient fluxes, denitrification and DNRA in a ameliorate N-effluxes from the sediment via the photoassimilation sub-tropical estuarine sediment. Hydrobiologia 628, 95e109. of dissolved N-species. 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