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

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Estuarine, Coastal and Shelf Science

journal homepage: www.elsevier .com/locate/ecss

Benthic metabolism and nitrogen dynamics in an urbanised tidalcreek: Domination of DNRA over denitrification as a nitrate reductionpathway

Ryan J.K. Dunn a,*, David Robertson b,1, Peter R. Teasdale b, Nathan J. Waltham c,2,David T. Welsh b

aGriffith School of Engineering, Gold Coast Campus, Griffith University, QLD 4222, Australiab Environmental Futures Centre, Gold Coast Campus, Griffith University, QLD 4222, AustraliacCatchment Management Unit, Gold Coast City Council, PMB 5042 Gold Coast Mail Centre, QLD 9729, Australia

a r t i c l e i n f o

Article history:Received 13 March 2013Accepted 26 June 2013Available online 9 July 2013

Keywords:benthic metabolismstormwater impactsnutrient fluxesdenitrificationdissimilatory nitrate reduction toammoniumsub-tropicalSaltwater Creek

* Corresponding author. Present address: Asia-PacifiSurfers Paradise, QLD 4217, Australia.

E-mail addresses: rdunn@apasa.com.au, ryanjkdun1 Present address: Science Museum (London), Exhib

United Kingdom.2 Present address: Centre for Tropical Water and

(TropWATER), James Cook University, QLD 4811, Aust

0272-7714/$ e see front matter � 2013 Elsevier Ltd.http://dx.doi.org/10.1016/j.ecss.2013.06.027

a b s t r a c t

Benthic oxygen and nutrient fluxes and nitrate reduction rates were determined seasonally under lightand dark conditions at three sites in a micro-tidal creek within an urbanised catchment (Saltwater Creek,Australia). It was hypothesized that stormwater inputs of organic matter and inorganic nitrogen wouldstimulate rates of benthic metabolism and nutrient recycling and preferentially stimulate dissimilatorynitrate reduction to ammonium (DNRA) over denitrification as a pathway for nitrate reduction. Storm-waters greatly influenced water column dissolved inorganic nitrogen (DIN) and suspended solids con-centrations with values following a large rainfall event being 5e20-fold greater than during thepreceding dry period. Seasonally, maximum and minimum water column total dissolved nitrogen (TDN)and DIN concentrations occurred in the summer (wet) and winter (dry) seasons. Creek sediments werehighly heterotrophic throughout the year, and strong sinks for oxygen, and large sources of dissolvedorganic and inorganic nitrogen during both light and dark incubations, although micro-phytobenthos(MPB) significantly decreased oxygen consumption and N-effluxes during light incubations due tophotosynthetic oxygen production and photoassimilation of nutrients. Benthic denitrification ratesranged from 3.5 to 17.7 mmol N m2 h�1, denitrification efficiencies were low (<1e15%) and denitrificationwas 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 Creeksediments would tend to increase rather than reduce dissolved nutrient loads to the downstream GoldCoast Broadwater and Moreton Bay systems. This may be especially true during wet periods whenincreased inputs of particulate organic nitrogen (PON) and suspended solids could respectively enhancerates of N-regeneration and decrease light availability to MPB, reducing their capacity to ameliorate N-effluxes through photoassimilation.

� 2013 Elsevier Ltd. All rights reserved.

1. Introduction

The biogeochemistry of coastal waterways is influenced bymanycompeting and interacting physical and biological factors (e.g.Rysgaard et al., 1995; Sundbäck et al., 2000; Bartoli et al., 2000;Azzoni et al., 2001; Sakamaki et al., 2006; Dunn et al., 2009;

c ASA Pty. Ltd., P.O. Box 1679,

n@hotmail.com (R.J.K. Dunn).ition Road, London SW7 2DD,

Aquatic Ecosystem Researchralia.

All rights reserved.

Spilmont et al., 2011; Pagès et al., 2012; Dunn et al., 2012a), result-ing in significant spatial and temporal variations inbenthic processes(Sundbäcketal., 2000;Welshetal., 2000;WilsonandBrennan,2004;Thornton et al., 2007; Nizzoli et al., 2007; Dunn et al., 2012a). Surfacesediments play a significant role in the microbially mediated trans-formations of nitrogen (Rysgaard et al., 1993, 1995; Fenchel et al.,1998; Dunn et al., 2012a) and therefore understanding the pro-cesses that influence the attenuation and recycling of N-species is ofgreat importance for coastal managers challenged with managingcompeting conservation and development land uses (Laima et al.,2002; Wilson and Brennan, 2004).

Coastal waterways are often subject to extensive urbanisation(Pauchard et al., 2006; Lee et al., 2006), which typically results inmajor changes in both the volume and quality of stormwater runoff

R.J.K. Dunn et al. / Estuarine, Coastal and Shelf Science 131 (2013) 271e281272

(Line and White, 2007; Bratieres et al., 2008). Urban stormwater isknown to transport elevated concentrations of heavy metals(Marsalek and Marsalek, 1997; Dunn et al., 2007a; Waltham et al.,2011), organic compounds (Mermillod-Blondin et al., 2005), sedi-ments (Brezonik and Stadelmann, 2002) and nutrients (Lee et al.,2006; Avila-Foucat et al., 2009) into receiving waters, usually un-treated and unfiltered. These anthropogenic inputs influencebenthic respiration rates and the nutrient status of receiving wa-terways due to the increased delivery of organic matter, and dis-solved and particulate nutrients. In some instances, increasednutrient loads can ultimately result in eutrophication of thereceiving water body (Nixon, 1995; Taylor et al., 2005). Alterationsin concentration gradients between the water column and surfacesediments induce change in the fluxes of oxygen and nutrientsacross the sediment-water interface, which can also influence ratesof nitrate reduction processes within the sediment (Blackburn andBlackburn, 1993; Fenchel et al., 1998). However, the effects of urbanrunoff on receiving water quality are highly site specific (US EPA,1983), making it difficult to predict the impacts or design appro-priate management and control programs (Brezonik andStadelmann, 2002).

Nutrient exchanges and nitrogen cycling pathways in shallowcoastal systems are generally quantified by scaling up individualmeasurements (Eyre and Ferguson, 2005 and references therein),which are often used as inputs within system models to determineallocations of management resources (Eyre and Ferguson, 2005).Therefore, a good understanding of the spatial and temporal vari-ability of nutrient dynamics is critical if these scaled up rates are togive reliable system-wide estimates. Knowledge of spatial andtemporal variations provide insight into influential factors con-trolling and maintaining benthic exchanges and nutrient cyclingpathways (Eyre and Ferguson, 2005). This is important for effectiveplanning and integrated management tools, and allows improvedpredictions regarding environmental changes due to anthropo-genic impacts (Sakamaki et al., 2006).

Saltwater Creek is a subtropical, urbanised creek within thesouthern Moreton Bay catchment (Queensland) and the estuary ispart of the Moreton Bay Marine Park. In 2005, Gold Coast CityCouncil (local government authority) completed an environmental

Fig. 1. Map (a) Australia, showing (b) southern Moreton Bay, and (c) sample sites within Saldischarge point).

inventory of Saltwater Creek (Tomlinson et al., 2006) in response topublic enquires regarding the health and ecology of the catchment.The study revealed considerable stress was being placed on thesystem due to elevated sediment and nutrient loads from storm-water inputs, and concluded that future planned urban develop-ment in the catchment would exacerbate this pressure, and furthercompromise public values and amenity of the catchment.

The objective of this study was to present an initial assessmentof the spatial and seasonal variations in benthic metabolism,nutrient fluxes, and rates of N-cycling processes in the intertidalsediments of Saltwater Creek and the impacts of stormwaters onwater column nutrients, N-cycling processes and physico-chemicalparameters. This urban catchment is representative of small coastalcatchments throughout Moreton Bay, Australia. The study is a blueprint for similar system based nutrient process investigations, andaims to provide the information necessary for the future manage-ment of the waterway.

2. Methods

2.1. Study location

Saltwater Creek is a micro-tidal estuarine creek with anurbanised and modified freshwater catchment located withinsouthern Moreton Bay (Australia). The creek system is approxi-mately 17 km long, flowing from its headwaters in Nerang StateForest to the Coombabah Creek estuary confluence which connectsto the Coomera River and the Gold Coast Broadwater (Fig. 1). Nat-ural vegetation is sparse along the estuary, and where present isdominated by mangroves (Avicennia marina, Rhizophora stylosa andAegiceras corniculatum) with saltmarsh (Sporobolus virginicus)located downstream towards the entrance of the creek. The upperestuary has several tidally connected residential canal estates(Benfer et al., 2007; Waltham and Connolly, 2011) and is revettedwith rock gabion foreshore armour for protection against erosion.The Helensvale Wastewater Treatment Plant also dischargedlicensed treated (secondary) wastewater to the upper estuary, onthe ebbing tide, until 2000 when it was decommissioned(Tomlinson et al., 2006). The intertidal zone of the creek is a

twater Creek (: represents the location of the Helensvale Wastewater Treatment Plant

R.J.K. Dunn et al. / Estuarine, Coastal and Shelf Science 131 (2013) 271e281 273

typically turbid system (Waltham, 2002; Tomlinson et al., 2006)and extends w10 km upstream from the creek mouth with a tidalamplitude of <1 m. Maximumwater depth is 4 m and decreases inthe upstream direction. The creek catchment is 30 km2 and hasrapidly changed from a forested environment to an extensive urbanresidential landscape, including golf course, canal estates, roads,and commercial and industrial estates (Waltham, 2002). The creekexperiences a wide range of hydrodynamic conditions and sedi-ment loadings according to inflow conditions (Webster andLemckert, 2002), complicated by interactions with the CoomeraRiver system (Webster and Lemckert, 2002). The catchmentdrainage network consists of overland and underground concretepipes that deliver untreated stormwater directly to the creek. Thelocal climate is characterised by warm/hot humid summers influ-enced by monsoonal trade winds. Thunderstorms are commonduring the summer period (NovembereJanuary) often resulting inintense short periods of catchment freshwater flow. In contrast,winters are typically characterised by low rainfall (average 20 mmper month; Waltham, 2002). Daily rainfall values for the Gold Coastregion, measured at Gold Coast Seaway (Fig. 1) by the AustralianBureau of Meteorology during the study year (2008) are shown inFig. 2.

2.2. Study design

Three sites located approximately 2.0, 6.0 and 9.5 km upstreamof the creek mouth (Fig. 1) were sampled in early and late Autumn(March and May), winter (July) and summer (December) of 2008.During each survey, six sediment cores were collected at each sitefor the determination of sediment-water column fluxes of oxygen,dissolved inorganic nitrogen (DIN; NOx (SNO�

2 þ NO�3 ) and NH�

4 ),total dissolved nitrogen (TDN) and dissolved organic nitrogen(DON), and rates of NO�

3 reduction processes. Cores collected forsediment incubations were collected from the three sites duringeach seasonal event over a one week period. Water samples werecollected, during each seasonal sediment collection period, foranalysis and core incubation during the high tide period immedi-ately following seasonal sediment collections. Additional watersamples were collected at each site, at high and low water, on the12th May 2008 following a two week dry period and the 3rd June2008 following a large rainfall event (90mmover previous 72 h and65mm in preceding 24 h period). Lastly, during the summer sampleperiod, three additional sediment cores were collected at each sitefor the characterisation of physico-chemical surface sedimentparameters.

Fig. 2. Daily recorded rainfall rates for the Gold Coast Seaway 2008 (data sourced:Australian Bureau of Meteorology).

2.3. Sediment and water collection

Sediment cores w15 cm in depth, for flux and process rate de-terminations, were hand collected at low water using plexiglass(200 mm diameter � 330 mm) tubes before transportation (within2 h) to the laboratory. Water used for core maintenance and incu-bation was collected during the following flood tide using clean(washed with deionised water [Milli-Q Element] and rinsed withcreek water) 40 L plastic containers. Triplicate water samples werecollected for the determination of DIN, TDN, chlorophyll-a (chl-a)and total suspended solids (TSS) concentrations. Surface sedimentlight intensity at each site was measured using an LI-COR radiationsensor (SA: LI-192SA quantum sensor). Light intensities weremeasured w15 cm above the sediment surface at a fixed moni-toring period (i.e. w12:00 � 1 h). In situ water measurements werecollected using a calibrated multi-probe monitoring unit (TPS 90-FLMV, TPS).

Cores collected for the determination of sediment characteris-tics (wet-bulk density, porosity, organic matter determined as loss-on-ignition (LOI550), grain size distribution, chl-a, phaeopigmentand bioavailable ammonium (NH4

þbio; porewater þ exchangeable

NHþ4 ) concentrations were hand collected using 50 mm

diameter � 400 mm PVC sample tubes. Cores were immediatelysliced into six depth horizons (0e1, 1e2, 2e4, 4e6, 6e10 and 10e15 cm) and stored in the dark (<4 �C) before freezing (�20 �C)within 2 h of collection.

2.4. Determination of oxygen and nutrient fluxes

Following collection, cores were immediately returned to thelaboratory, carefully filled with creek water and submerged inholding tubs (220 L) at seasonally measured in situ water temper-atures (23, 22, 19 and 24 � 1 �C in autumn (March), autumn (May),winter and summer, respectively). An aquarium pump and air-stone was fitted within each core to facilitate water circulationand aeration and triplicate cores were equilibrated under light anddark conditions for w12 h, at the seasonally measured in situ lightintensity (w80 mE m�2 s�1). Following equilibration, the light anddark conditions were reversed, w40% of the water in the holdingtanks replaced with fresh site water and the cores re-equilibratedunder the new conditions for 2 h.

To initiate incubations, the air-stones were removed from thecores with the water lowered to below the core rims. The aquariumpumps were left within each core to maintain water circulation(resuspension of core sediments was avoided). Initial water sam-ples for dissolved oxygen, DIN and TDN were then taken and thecores were then closed using floating plastic lids to prevent gaseousexchange with the atmosphere. Cores were incubated for w1.5 hand at the end of this period the aquarium pumps were stopped,the floating lids removed and final timewater samples immediatelycollected. Flux rates (mmol m�2 h�1) were calculated from thechange in water column concentrations of the individual solutesfollowing Welsh et al. (2000).

2.5. Determination of nitrate reduction rates

Following flux incubations, the air-stones and aquarium pumpswere replaced in each core, the water level in the tanks was raisedto above the core tops and the cores were left to equilibrate forw2.5 h before the determination of NO�

3 -reduction rates using theisotope paring technique (IPT), as modified for simultaneousdetermination of denitrification and dissimilatory nitrate reductionto ammonium (DNRA). IPT core treatment and sample collectionfollowed the protocol described by Dunn et al. (2012a).

R.J.K. Dunn et al. / Estuarine, Coastal and Shelf Science 131 (2013) 271e281274

Rates of total denitrification (D14), coupled nitrification-denitrification (DN) and denitrification of NO�

3 from the overlyingwater (DW) were determined following Nielsen (1992). DNRA ratesbased on water column NO�

3 (DNRAW) were calculated from the15N-enrichment of the water column NO�

3 pool and the extractedsediment bioavailable NHþ

4 pool (Risgaard-Petersen and Rysgaard,1995). DNRA rates coupled to nitrification (DNRAN) were esti-mated from the rate of DNRAW and the ratio between DN and DW(Risgaard-Petersen and Rysgaard, 1995) and total rates of DNRA asthe sum of DNRAW and DNRAN. Anammox is recognised as aninterference when using IPT that can lead to overestimates ofdenitrification rates, as it also generates labelled N2 speciesfollowing 15NO�

3 additions (Risgaard-Petersen et al., 2003). How-ever, in shallow water sediments anammox is a minor source of N2compared to denitrification (Dalsgaard et al., 2005; Burgin andHamilton, 2007), especially in tropical systems (Dong et al., 2011).Therefore, the authors believe that reported estimates of denitri-fication are valid, although it should be noted that the term deni-trification here also includes a small portion of N2 produced viaanammox.

2.6. Sample handling and analytical techniques

Sediment wet-bulk density, porosity and LOI550 were deter-mined following Dunn et al., 2007b. The proportion of clay and silt(<63 mm), sand (180 < x > 63 mm) and gravel (>180 mm) sedimentfractions were determined by dry sieving and are expressed as apercent dry weight. Sediment chl-a and phaeopigment concen-trations were determined following freeze drying and acetoneextraction of 1 cm3 aliquots of surface sediment following Lorenzen(1967). Sediment NHþ

4 bio concentrations were determinedfollowing extraction of 1 cm3 of homogenised sediment for 24 h in9 ml 2 M KCl (Nizzoli et al., 2005).

Water samples were collected using acid washed (10% v/v HCl),sample rinsed, low density polyethylene sample bottles (Nalgene).Dissolved inorganic and total nutrient samples were immediatelyfiltered through pre-washed, pre-ashed 25 mm GF/F filters andstored frozen (�20 �C) until analysed. Chl-a samples were filteredthrough 25 mm pre-washed, pre-ashed GF/C filters immediatelyfollowing water collection and the filters stored frozen until ana-lysed (APHA, 2005). TSS samples were filtered through pre-washed,pre-ashed 47 mm GF/F filters. During sediment incubation, watersamples were collected using acid washed and Milli-Q elementwater rinsed 60 ml plastic syringes and tubing. Nutrient sampleswere filtered (GF/F) and stored frozen (�20 �C) until analysis. Dis-solved oxygen samples were carefully transferred to gas-tight 12mlglass vials (Exetainer, Labco Ltd.), fixed with 100 ml of manganoussulfate and alkali-iodide-azide solution (APHA, 2005), stored at 4 �Cand analysed within 48 h according to theWinkler titration (Azide-modification method, APHA, 2005).

DIN concentrations were directly determined by an automatednutrient analyser (Easychem Plus Random Access analyzer, SysteaAnalytical Technologies). TDN concentrations were determined asNOx (NO�

3 þ NO�2 ) following digestion (potassium persulfate/

Table 1Mean (�SD) depth integrated (0e15 cm) sediment characteristics and surface (0e1 cm)particle size distributions, density, porosity, LOI550 and NHþ

4 bio data and n ¼ 3 for Chl-a

Site Particle size distribution (%) Density Porosity

<63 mm 63e180 mm >180 mm (g cm�3)

1 30 � 9.4 60.2 � 5.1 11.9 � 5.4 1.53 � 0.13 69.0 � 4.42 56.4 � 7.5 31.9 � 7.2 11.5 � 3.4 1.80 � 0.18 61.3 � 7.13 43.0 � 12.9 40.1 � 9.8 16.6 � 7.6 1.55 � 0.17 52.7 � 6.4

sodium hydroxide digestion solution; 2 � 45 min digestions;120 �C). Dissolved organic nitrogen (DON) was determined by dif-ference (DON ¼ TDN � (NOx þ NHþ

4 ). Milli-Q Element water andfiltered low nutrient seawater were used for all sample preparationand analyses. Natural filtered seawater references standards andinternal standards were used for quality assurance. Digestion effi-ciency was established through the analysis of digested seawaterreference standards. Recoveries of all nutrients from the certifiedstandards ranged from 80 to 109%. Dissolved N2 concentrations andthe proportions of 29N2 and 30N2 and 15N enrichment of sediment-NHþ

4 bio pools were analysed at the National EnvironmentalResearch Institute (Silkeborg), Denmark.

2.7. Macrofauna dynamics

Following determinations of nitrate reduction rates, sedimentsfrom each incubation core plus the corresponding KCl-sedimentslurry were pooled and sieved (250 mm mesh) to recover burrow-ing macrofauna. Macrofauna were rinsed with freshwater andpreserved in 70% ethanol, counted and identified.

2.8. Statistical analyses

Variation in flux and nitrate reduction rates were analysed usingthree-way analysis of variance (ANOVA) with light/dark conditions,month (season) and site fixed, and interaction included. Seasonwasa fixed factor here as each survey represents a single point in time,with the timing of each survey representative of known maximaldifferences in local environmental conditions (i.e. temperature,rainfall, flow). Prior to analyses the assumption of homogeneity ofvariances was tested using box and whisker plots, before and aftertransformation (Quinn and Keough, 2008). Variances were beststabilised with a log (x) transformation. Post hoc comparisons wereperformed using Tukey’s HSD. Correlations between physico-chemical sedimentary conditions, faunal communities, solutefluxes and nitrate reduction rates were analysed using Pearsoncorrelation analysis (2-tailed). Statistical significance was deter-mined at a ¼ 0.05. All statistical analyses were performed usingSPSS Windows (SPSS Inc., version 19).

3. Results

3.1. Sediment characteristics

Sediment characteristics were relatively homogenous over the0e15 cm depth horizon sampled and are therefore expressed asdepth integrated means, with the exception of Chl-a and phaeo-pigment concentrations, which were only determined in the sur-face 1 cm (Table 1). Sediments at all sampling sites were dominatedby the <63 mm and 63e180 mm size fractions, which togetheraccounted for 83e90% of the total particle size distribution. Sedi-ment organic matter content (LOI550), increased from 5.8� 1.9% drywt at site 1 (closest to creek mouth) to 8.0 � 2.2 at site 3 (mostupstream estuary site). Sediment NHþ

4 bio concentrations ranged

concentrations of Chl-a and phaeopigment concentrations in the estuary. n ¼ 18 forand phaeopigment content.

LOI550 NH4þbio Chl-a Phaeopigment

(nmol g dry wt�1) (mg g dry wt�1) (mg g dry wt�1)

8.0 � 0.4 238 � 145 1.5 � 0.4 2.1 � 0.56.2 � 1.6 201 � 73 0.6 � 0.1 0.8 � 0.15.8 � 1.9 198 � 102 0.9 � 0.8 1.3 � 1.1

Table 3Water column nutrient concentrations and physico-chemical parameters at eachsampling site at high and low tide during dry conditions (12th May 2008; no rainfallin preceding 2 weeks) and wet conditions following a significant rainfall event (3rdJune 2008; 90 mm over 72 h, 65 mm in preceding 24 h). PN represents particulatenitrogen, all other abbreviations as in Table 3.

Sample site,conditionsand tide

NOx

(mM)NHþ

4(mM)

DIN(mM)

DON(mM)

TDN(mM)

PN(mM)

TSS(mg L�1)

DrySite 1 high 0.6 0.5 1.1 28.2 29.1 31.3 19.5Low 0.6 0.4 0.9 28.2 29.1 31.3 23.5Site 2 high 0.6 0.8 1.4 25.4 26.8 28.8 30.5Low 0.5 0.8 1.3 29.7 32.9 35.4 29.4Site 3 high 0.6 2.2 3.1 30.8 33.7 35.4 25.5Low 1.0 2.6 3.8 29.1 32.9 36.3 28.5

WetSite 1 high 18.1 5.4 25.7 15.8 41.5 44.7 97.0Low 19.9 5.3 24.1 17.2 41.4 44.5 83.0Site 2 high 31.9 3.3 36.0 19.1 55.1 59.2 263.0Low 34.1 3.2 38.1 4.0 42.1 50.6 250.0Site 3 high 33.4 6.3 40.3 6.1 46.4 58.4 303.3Low 27.3 7.4 35.5 11.6 47.1 57.5 283.3

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,respectively. Chl-a and phaeopigment concentrations were greatestat the seaward site 1.

3.2. Temporal variations in water column characteristics

Seasonal water column DIN concentrations ranged between1.54 � 0.68 and 10.33 � 0.88 mM (Table 2). Maximum DIN con-centrations were measured during the summer wet season, coin-ciding with reduced salinity. DIN concentrations were typicallydominated by the contribution of NOx, which represented 63.1% ofthe total DIN pool. However, DON was the dominant dissolved ni-trogen species, representing 89.4e98.2% of TDN, with the lowestcontribution of DON to the TDN pool recorded during the summersurvey, despite the highest measured DON concentration alsooccurring in summer.

Dry and wet event sampling showed that following a significantrainfall event in June (90 mm over 72 h, 65 mm in preceding 24 h),concentrations of DIN, NHþ

4 and NOx were 5e25-fold greater thanthose in May or July, with the highest concentrations measured atthe upstream sites (Table 2 and 3). In contrast, DON concentrationsduring this wet event were 50% or less of those measured duringthe May or July surveys, with the lowest DON concentrationsmeasured at sites 2 and 3. Similarly, concentrations of total sus-pended solids and particulate nitrogen were also elevated byapproximately 2 and 10-fold respectively, under wet compared todry conditions, with this effect again most evident at the upstreamsampling sites (Table 3).

3.3. Burrowing macrofauna dynamics

In total 420 individuals were collected, consisting of three spe-cies of surface and sub-surface deposit and filter feeding macro-faunal groups: worm (Simplisetia aequisetis), amphipod(Victoriopisa australianesis), and crab (Heloecius cordiformis). Mac-rofauna densities ranged from 0 to 605, 0 to 446 and 0 to 64 in-dividuals m�2 (ind m�2) for worms, amphipods and crabs,respectively. The highest combined abundance occurred in summer(Fig. 3). V. australiensis dominated the surface sediments at sites 1and 2 accounting for 45 and 50% of the total macrofaunal density,respectively. In comparison, at site 3 the abundance of S. aequisetisaccounted for 56% of the total abundance.

3.4. Sediment-water column oxygen and dissolved inorganicnutrient fluxes

Sediments at all sites were consistent sinks for water columndissolved oxygen during both light and dark incubations (Fig. 4),although rates of oxygen consumption were significantly lowerduring light compared to dark incubations (Table 4). Dark oxygenfluxes varied from �2900 to �2100, �2200 to �1900 and �4200to�1700 mmol m�2 h�1 at sites 1, 2 and 3, respectively. Estimates ofgross community primary production calculated by difference

Table 2Seasonal variations in water column dissolved nitrogen concentrations and physicoAut ¼ Autumn, Win ¼winter, Sum ¼ summer, DIN ¼ dissolved inorganic nitrogen, DONTemp. ¼ temperature and TSS ¼ total suspended sediments.

Season Dissolved nutrients

NOx (mM) NHþ4 (mM) DIN (mM) DON (mM) TDN (mM

Aut (Mar) 1.87 � 0.13 1.67 � 0.50 3.54 � 0.37 44.0 � 3.2 47.2 � 3Aut (May) 0.81 � 0.44 0.74 � 0.45 1.54 � 068 55.2 � 4.4 56.2 � 3Win 4.68 � 0.30 0.57 � 0.34 5.25 � 0.31 44.8 � 5.6 48.6 � 4Sum 7.23 � 0.31 5.25 � 0.31 10.33 � 0.88 69.1 � 9.5 77.3 � 7

between light and dark oxygen fluxes ranged between 400 (site 3,autumn;May) to 2000 mmolm�2 h�1 (site 1, autumn;May). Rates ofgross production were highest at site 1 and lowest at site 3, exceptduring summer, where the highest rate occurred at site 3.

Sediments at all sites in all seasons were a net source of NHþ4

during dark incubations and sediment dark NHþ4 effluxes (Fig. 5)

and oxygen demands were significantly correlated (r ¼ 0.682,p < 0.001). During light incubations, sediments were either smallsinks for NHþ

4 or showed much reduced rates of efflux. SeasonalNHþ

4 fluxes showed significant seasonal trends with highest ef-fluxes measured during summer and also significant differencesbetween sampling sites with the sediments of sites 2 and 3 beinggreater sources of NHþ

4 compared to site 1. Comparable to NHþ4

fluxes, sediments were also sources of NOx under dark conditionsand small sinks or greatly reduced sources of NOx under lightconditions (Fig. 5). NOx fluxes also demonstrated significant sea-sonal differences (Table 4) with highest effluxes measured insummer and the lowest in the autumn and winter periods. As bothNHþ

4 and NOx showed similar spatial and temporal trends, overallDIN fluxes followed the same pattern with sediments being largesources of DIN under dark conditions, slight sinks or greatlyreduced sources of DIN under light conditions with the highest DINeffluxes occurring in summer and the lowest in winter (data notshown). However, fluxes of DIN were small compared to DON,which represented 50e69% of the annual TDN efflux. Sedimentswere significant sources of DON in all seasons, at all sites duringboth light and dark incubations. Maximum DON effluxes occurredin winter (July) and lowest in autumn (March), although themaximum mean contribution of DON to dark TDN effluxes (87%)occurred in May, with the minimum contribution (42%) measuredin summer (December). As a result of the constant efflux of DON,

-chemical parameters. All data are mean values (�SD) (n ¼ 3). Abbreviations:¼ dissolved organic nitrogen, TDN ¼ total dissolved nitrogen, chl-a ¼ chlorophyll-a,

Physico-chemical parameters

) Chl-a (mg L�1) pH Temp. (oC) Salinity (&) TSS (mg L�1)

.9 3.10 � 0.31 8.1 � 0.1 24.1 � 0.2 28.4 � 0.3 28.3 � 2.1

.9 2.78 � 1.10 8.2 � 0.1 22.3 � 0.1 27.8 � 0.2 30.8 � 4.2

.2 2.92 � 0.49 8.0 � 0.1 20.4 � 0.3 28.2 � 0.3 26.3 � 2.7

.4 4.01 � 0.74 7.8 � 0.2 26.3 � 0.1 25.2 � 0.2 33.6 � 5.1

Fig. 3. Mean (�SD) a) seasonal abundance, and b) annual site abundance of macro-fauna species across all sample cores (n ¼ 72). Species are Heloecius cordiformis (crab),Victoriopisa australianesis (amphipod) and Simplisetia aequisetis (worm).

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

R.J.K. Dunn et al. / Estuarine, Coastal and Shelf Science 131 (2013) 271e281276

sediments at all sites were net sources of TDN to the water columnin all seasons during both light and dark incubations. TDN effluxesshowed significant (p < 0.01) differences between light and darkconditions with larger TDN effluxes occurring under dark condi-tions (Table 4 and Fig. 5), although this was largely due to the DINcomponent.

3.5. Nitrate reduction processes

Total nitrate reduction rates varied between 9.2 mmol-Nm�2 h�1

(site 1 autumn, dark incubation) to 110.6 mmol-N m�2 h�1 (site 2summer, light incubation) (data not shown) with a significantseasonal trend apparent with the highest rates measured in sum-mer (Table 4). DNRA was the dominant pathway for nitratereduction at all sampling sites in all seasons (Figs. 6 and 7), andaccounted for 64.9 to 81.8 and 70.1 and 86.0%, respectively, of lightand dark total nitrate reduction rates. The maximum contributionof DNRA to total nitrate reduction was measured during the sum-mer survey when DNRA accounted for on average 80.3 and 86.0% oftotal light and dark nitrate reduction, respectively, across thesample sites. Total denitrification rates ranged between3.5 � 0.5 mmol-N m�2 h�1 to 17.7 � 5.5 mmol-N m�2 h�1 (Fig. 6),with significant (p < 0.001) differences measured between lightand dark conditions. Mean total DNRA ranged between5.1 � 6.2 mmol-N m�2 h�1 to 92.3 � 94.6 mmol-N m�2 h�1 (Fig. 7),with an apparent significant seasonal trend with highest ratesrecorded in summer (Table 4).

Nitrification was the principal source of nitrate fuelling nitratereduction processes in the sediment (Figs. 6 and 7), with DN andDNRAN accounting for 81.2 � 14.6% and 74.4 � 20.9% of averageannual nitrate reduction under light and dark conditions, respec-tively. In contrast, rates of DW and DNRAW were of lower impor-tance, and water column nitrate was only a significant source ofnitrate for denitrification and DNRA within the creek sediments insummer (Figs. 6 and 7), which corresponded with higher NOxconcentrations in the water column (Table 2). Sediment denitrifi-cation efficiencies calculated as N2eN/(N2eN þ DIN efflux) � 100(Eyre and Ferguson, 2002) ranged between 0.72 and 79.1% with anoverall combinedmean of 14.9� 17.3% across all sampling sites andseasons. Lower efficiencies generally corresponded with highersediment oxygen demands and greater efficiencies were observedwith lower sediment oxygen demands. Fig. 8 illustrates meandenitrification efficiencies for individual incubation cores at sedi-ment oxygen demands of<1000, 2000< x> 1000 and>2000 mmolO2 m2 h�1, respectively.

4. Discussion

4.1. Water column nutrient concentrations

The seasonal pattern for water column DIN concentrations wascomparable with those recorded in previous studies withinSouthern Moreton Bay (Tomlinson et al., 2006; Dunn et al., 2007c,2012a; Eyre et al., 2010a, 2012b). Elevated concentrations coincidedwith summer wet season conditions characterised by greaterexternal inputs of stormwater. DON was the dominant componentof the water column TDN pool and also dominated sediment-water

Table 4Summary of results of three-way ANOVA for seasonal solute flux and nitrate reduction rate measurements among sample sites under light and dark conditions. Significantoutcomes (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 thesediment; all other abbreviations as in legend to Table 3.

DF Sediment-water column fluxes Nitrate reduction rates

O2 NHþ4 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 0.549Month 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 0.028Site 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 0.249L � 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 0.312L � S 2 0.008 0.290 0.998 0.840 0.684 0.876 0.449 0.213 0.755 0.219 0.088 0.135 0.090M � 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 0.315L � 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

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column nutrient effluxes to a similar extent, indicating that underdry conditions the composition of the water column TDN pool wasdriven by nutrient cycling processes in the sediment.

In contrast, following a significant rainfall event in June 2008,DIN concentrations in the creek waters increased dramatically byup to 25-fold, to values that are the highest recorded concentra-tions for SouthernMoreton Bay (Tomlinson et al., 2006; Dunn et al.,2007c, 2012a; Eyre et al., 2010a, 2012b). This increase in DIN con-centrations was accompanied by substantial increases in TSS andparticulate nitrogen concentrations, and reduced salinity and DONconcentrations. These dramatic changes, clearly demonstrate thatduring this wet period, stormwater run-off rich in DIN, suspendedsolids and particulate nitrogen was a major influence on overallphysico-chemical conditions in the creek waters. Similar patternshave been observed elsewhere (e.g. Goonetilleke et al., 2005; Flintand Davis, 2007; Beck and Birch, 2012; Sood et al., 2012) supportinga model that the physico-chemical conditions in the water column

Fig. 5. Seasonal sediment-water column fluxes of nitrate þ nitrite (NOx), ammoniuNOx þ NHþ

4 þ DON) at each sample site during light and dark conditions. Negative values indof the solute by the sediment (efflux). Data are mean values and error bars indicate the sta

can be driven by the volume of stormwater flow emanating fromthe catchment. Additionally, given that in Southern Moreton Bay,w30% of the annual regional rainfall falls between December andFebruary (summer) and rainfall events much larger than thestudied June event are also common around this period (e.g. eventsin late November and early January of the study year; Fig. 2), thepotential influence of stormwaters on physico-chemical conditionsin Saltwater Creek could conceivably be even greater than thosemeasured here, which poses greater challengers for managers.

Nutrient inputs associated with stormwaters following heavyrainfall have been linked to increased phytoplankton and algalgrowth in the nearby Coomera and Pimpama River estuaries (Eyreet al., 2010b). However, given the relatively high TSS concentrationswhich persisted throughout the year in Saltwater Creek and thevery high TSS concentrations associated with stormwater inputs, itis likely that primary production in the water column is lightlimited and that benthic production is essentially limited to shallow

m (NHþ4 ), dissolved organic nitrogen (DON) and total dissolved nitrogen (TDN;

icate consumption of the solute by the sediment (influx) and positive values productionndard deviation of the mean flux (n ¼ 3).

Fig. 6. Seasonal rates of denitrification based on nitrate diffusing from the watercolumn (DW), denitrification coupled to nitrification (DN), and total denitrification rates(D14) at each sample site during light and dark conditions. Data are mean values anderror bars indicate the standard deviation of the mean rate (n ¼ 3).

Fig. 7. Seasonal rates of DNRA based on nitrate diffusing from the water column(DNRAW), DNRA coupled to nitrification (DNRAN), and total DNRA (DNRA14) at eachsample site during light and dark conditions. Data are mean values and error barsindicate the standard deviation of the mean rate (n ¼ 3).

Fig. 8. Mean denitrification efficiencies for individual incubation cores at categorisedsediment oxygen demands of<1000 mmol O2 m2 h�1 (n ¼ 6), 2000< x > 1000 mmol O2

m2 h�1 (n ¼ 29) and >2000 mmol O2 m2 h�1 (n ¼ 13). Error bars indicate the standarddeviation of the mean denitrification efficiencies.

R.J.K. Dunn et al. / Estuarine, Coastal and Shelf Science 131 (2013) 271e281278

subtidal and intertidal zones during periods of low water. Conse-quently, the bulk of the DIN and suspended solids entering thecreek in the stormwaters are likely to be exported to the down-stream Gold Coast Broadwater, where TSS may negatively impacton seagrass communities (McClennan and Sumpton, 2005; Pagèset al., 2012).

4.2. Benthic metabolism and nutrient fluxes

The sediments throughout the creek system were a sink forwater column dissolved oxygen under both light and dark condi-tions, as has also been reported in intertidal areas of other eastcoast Australian estuaries (Eyre and Ferguson, 2002, 2005; Qu et al.,2006; Dunn et al., 2012a). These oxygen dynamics are typical ofsediments receiving high organic matter loads from externalsources (organic matter inputs > local production ¼ net hetero-trophic conditions (Viaroli et al., 2004)) and may at least in part bedue to organic matter loading from urban stormwater. This sus-tained sediment oxygen demand during both light and dark con-ditions may negatively impact the water column, in particular,contributing to the low dissolved oxygen saturation valuesmeasured in this creek system (Waltham, 2002; Tomlinson et al.,2006). Sustained sediment oxygen demand and reduced watercolumn oxygen concentrations would also favour anaerobic

R.J.K. Dunn et al. / Estuarine, Coastal and Shelf Science 131 (2013) 271e281 279

microbial metabolisms and the accumulation of the toxic reducedmetabolites (e.g. sulfides and Fe(II)) within the sediment. The dis-tribution of these compounds, especially sulfides, is a majordeterminant of benthic fauna distributions and could be respon-sible for the low abundance, biomass and biodiversity of infaunaobserved in the creek sediments (Cosser, 1989).

Micro-phytobenthos were present at the sediment surface, asindicated by chl-a concentrations and shifts in sediment oxygenfluxes measured during light compared to dark conditions,although oxygen fluxes remained negative during light conditions.The consistent consumption of oxygen by the sediment during lightincubations reflects the high inputs of organic matter, which driverespiratory processes and the high turbidity of the water column.Increased turbidity limits light availability for photosynthesis and itis likely that in these intertidal sediments microalgal production isessentially limited to the low water period, during sedimentemersion. However, this production would not contribute to re-oxygenation of the water column. It should be noted that datafrom our light incubated cores represent a best case scenario, as thebulk of the sub-tidal creek sediments are unlikely to receive suffi-cient light to support microalgal communities and would thereforeact as strong constant sinks for oxygen regardless of prevailing lightconditions at the water surface.

Heterotrophic metabolism in the sediments drove sustainedeffluxes of dissolved nitrogen species during all seasons at all sites,although photoassimilation of N did significantly reduce effluxesduring light incubations. Highest DIN and TDN effluxes occurredduring summer and effluxes showed similar magnitudes and sea-sonal dynamics to those reported for other turbid east coastAustralian estuaries (Qu et al., 2005; Eyre et al., 2010a; Dunn et al.,2012a). DON was the dominant contributor to sediment TDN ef-fluxes, accounting for on average 50e70% of the TDN efflux acrosssites. Model and mesocosm simulations (Blackburn and Blackburn,1993; Sloth et al., 1995) indicate that high contributions of DON tooverall TDN effluxes are characteristic of sediments that receiveelevated organic matter loads that are mineralised primarily at/ornear to the sediment-water interface and notmixed into the deepersediment layers by bioturbation. Thus, the high TDN effluxes andhigh contribution of DON to these effluxes may reflect the highorganic loads the sediments receive from stormwater inputs andthe limited abundance and biodiversity of macrofauna, whichlimits mixing of this organic matter into the deeper sedimentstrata. Consequently, organic matter is mineralised at or close to thesediment-water interface resulting in a high proportion of the earlymineralization products such as DON and NHþ

4 escaping to thewater column by diffusion before they can be further metabolized(Blackburn and Blackburn, 1993; Sloth et al., 1995).

4.3. Nitrate reduction processes

Denitrification rates in the Saltwater Creek sediments werecomparable to those measured in western Moreton Bay (w6ew27 mmol-N m�2 h�1; Ferguson et al., 2007), Coombabah Lake (<1to 6 mmol-N m�2 h�1; Dunn et al., 2012a) and the (regional scale)Brunswick estuary (w1e58 mmol-N m�2 h�1; Eyre and Ferguson,2005). Anammox was not directly measured, but would be ex-pected to make only a minor contribution to overall N2 productionin the studied estuary. For example, in a study of 40 sites in 9 es-tuaries in southern England, Nicholls and Trimmer (2009) foundthat in sediment slurries amended with 15N substrates, that ana-mox accounted for only 1e11% of total N2 production. In this study,the % contribution of anammox to N2 production increased withwater column nitrate concentration over the range of 3e790 mM.However, at the water column nitrate concentrations in the 1e7 mMrange measured in Saltwater Creek during our seasonal samplings,

the % contribution of anamox to N2 production was typically below1% (Nicholls and Trimmer, 2009). Therefore, we are confident thatour denitrification measurements provide a true reflection of N-loss as N2 from the Saltwater creek sediments.

The creek sediments also had low denitrification efficiencies,which were again comparable with estuaries elsewhere (e.g.Berelson et al., 1998; Heggie et al., 1999; Eyre and Ferguson, 2002;Cook et al., 2004; Dunn et al., 2012a), presumably as a result of highcarbon loadings. Large decreases in denitrification efficiency havebeen reported at carbon decomposition rates >w1500 mmol CO2m�2 h�1 (Eyre and Ferguson, 2002) and w1250 mmol CO2 m�2 h�1

(Berelson et al., 1998). Therefore, the high carbon loadings withinthe surface sediments of Saltwater Creek, as indicated by elevatedLOI550 values and oxygen demands, presumably contributed to theobserved low denitrification efficiencies. During this study thelowest benthic denitrification efficiencies measured in individualincubation cores coincided with those having the highest sedimentoxygen demands and denitrification efficiency decreased steeply incores with oxygen demands >2000 mmol m�2 h�1 (Fig. 8). Thishypothesis is supported by the significant increase in rates ofcoupled nitrification-denitrification that occurred during light in-cubations, which indicates that nitrification and hence denitrifi-cation rates in the Saltwater Creek sediments were limited byoxygen rather than NHþ

4 availability and that this limitation can beoffset by photosynthetic oxygen production, as has been observedin other strongly heterotrophic sediments colonised by benthicmicroalgae (An and Joye, 2001; Dunn et al., 2012a). However, thedominance of DNRA over denitrification as a sink for nitrate wouldalso substantially contribute to the low denitrification efficienciesin the sediments through competition between denitrification andDNRA for NOx, although DNRA itself is also favoured by highsediment carbon loadings and metabolic rates (Tiedje, 1988;Christensen et al., 2000; Nizzoli et al., 2006). Consequently, re-ductions in carbon loadings to the creek from stormwater wouldpresumably enhance the sediment denitrification efficiency andthe natural removal of nitrogenwithin the creek sediments, both byreducing sediment respiration rates and by favouring denitrifica-tion over DNRA as a pathway for nitrate reduction.

DNRA was the dominant pathway for nitrate reduction in allseasons at all sites, accounting for w75% of total nitrate reduction.Rates of total DNRA, DNRAW and DNRAN all showed the same sea-sonal pattern with highest rates occurring during summer. Theseresults are in agreement with those other recent studies of tropicaland sub-tropical estuaries (Dong et al., 2011; Dunn et al., 2012a;Molnar et al., 2013) and support the hypothesis that DNRA is avery much more important process in the N-dynamics of tropicaland sub-tropical compared to temperate estuarine sediments.Several factors have been proposed to favour DNRA over denitrifi-cation as a pathway for nitrate reduction including; high tempera-ture, high ratios of labile organic carbon toNO�

3 ratios, elevated ratesof benthicmetabolism, lowNO�

3 availability, and reduced, especiallysulphidic sediment conditions (see Nizzoli et al., 2006 and refer-ences therein), although the influence of the factors is difficult toseparate as they are often coincident (Nizzoli et al., 2006). Our dataare in general agreement with these regulatory mechanisms asSaltwater Creekwater temperatures were typically above 20 �C andbenthic respiration rateswere elevated throughout the year, and thelargest dominance of DNRA over denitrification occurred duringsummer when temperature and rates of benthic respiration weremaximal and the sediment redox conditions would likely be mostreduced.

Nitrification was the primary source of nitrate fuelling nitratereduction in the Saltwater Creek sediments with DN and DNRANaccounting forw75% of total nitrate reduction under both light anddark conditions. As discussed earlier, rates of DN were significantly

R.J.K. Dunn et al. / Estuarine, Coastal and Shelf Science 131 (2013) 271e281280

stimulated during light incubations indicating that microalgalphotosynthesis played a regulatory role and oxygen rather thanNH4

þ availability was limiting nitrification rates. Thus, despite theirrelatively low biomass benthic microalgae still played a major rolein regulating N-fluxes and N-cycling processes through the pho-toassimilation of N-species and the production of oxygen.

Although nitrification was the major source of nitrate for deni-trification, rates of denitrification dependent on nitrate diffusingfrom the overlying water column (DW) showed significant seasonalvariations, with maximum rates measured during summer coin-ciding with the maximum measured concentrations of NOx in thewater column. Such regulation of DW by water column nitrateconcentrationwould be expected and has frequently been observedin marine sediments (e.g. Rysgaard et al., 1995; Christensen et al.,2000; Nizzoli et al., 2006; Hietanen and Kuparinen, 2008), as therate of supply of nitrate from the water column to the sedimentdenitrification zone largely depends on the concentration differ-ence between the two zones. In Saltwater Creek this effect wouldbe amplified, as peak water column NOx concentrations occurred insummer when sediment oxygen demand was maximal and there-fore oxygen penetration into the sediment minimal (Fenchel et al.,1998). Such conditions would further enhance the concentrationgradient of nitrate and therefore the diffusion rate of nitrate fromthe water column to the sediment denitrification zone. It wouldalso be expected that, during wet periods, increased inputs of ni-trate to the creek during stormwater flow would greatly stimulaterates of DW, as following the sampled high rainfall event in June thewater column NOx concentrations were 4e34-fold higher thanthose measured during the seasonal sampling campaigns. How-ever, due to the dominance of DNRA as a pathway for nitratereduction in the creek sediments, this increase in DW would havelittle impact on the eutrophication status of the creek or thenutrient loads transported to downstream water bodies.

In terms of eutrophication, DN can be considered as the sedimentscapacity to eliminate internal nitrogen loads and DW as the sedi-ments capacity to eliminate external nitrate loads during transport(Nizzoli et al., 2006). Whereas, DNRAW represents an input of N tothe sediment compartment, as nitrate diffusing into the sediment isconverted to ammonium and DNRAN constitutes a futile cycle, asnitrate produced in the sediment from ammonium by nitrification issimply recycled back to ammonium in the sediment by DNRA(Nizzoli et al., 2006). Since over 75% of nitrate produced in thesediments by nitrification or diffusing into the sediment from theoverlying water columnwas converted to ammonium via DNRA, thecapacity of the sediments to eliminate internal or external N-loadswas extremely limited. Indeed, as benthic rates of total denitrifica-tionwere small compared to the benthic effluxes of DIN and TDN, thenet effect of N-cycling in the creek sediments would be to enrich theoverlying waters with dissolved N-species during transport. Thismay be especially true during wet periods for two principal reasons.Firstly, as stormwater increased particulate N concentrations in thewater column approximately 2-fold, sedimentation of these partic-ulates would be expected to further stimulate benthic metabolismand nutrient effluxes. Secondly, stormwater increased TSS concen-trations in the creekwaters 3e10-fold, whichwould further decreaselight availability to benthicmicroalgae and therefore their capacity toameliorate N-effluxes from the sediment via the photoassimilationof dissolved N-species.

Acknowledgments

This research was financially supported by the Gold Coast CityCouncil Catchment Management Unit in association with the

Griffith Centre for Coastal Management (Griffith University, GoldCoast Campus).

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