Links between contaminant hotspots in low flow estuarine systems and altered sediment biogeochemical processes

Accepted Manuscript Links between contaminant hotspots in low flow estuarine systems and altered sediment biogeochemical processes Michael Sutherland, Katherine A. Dafforn, Peter Scanes, Jaimie Potts, Stuart L. Simpson, Vivian X.-Y. Sim, Emma L. Johnston PII:

S0272-7714(16)30284-0

DOI:

10.1016/j.ecss.2016.08.029

Reference:

YECSS 5219

To appear in:

Estuarine, Coastal and Shelf Science

Received Date: 30 December 2015 Revised Date:

25 July 2016

Accepted Date: 27 August 2016

Please cite this article as: Sutherland, M., Dafforn, K.A., Scanes, P., Potts, J., Simpson, S.L., Sim, V.X.-Y., Johnston, E.L., Links between contaminant hotspots in low flow estuarine systems and altered sediment biogeochemical processes, Estuarine, Coastal and Shelf Science (2016), doi: 10.1016/ j.ecss.2016.08.029. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

AC C

EP

TE D

M AN U

SC

RI PT

ACCEPTED MANUSCRIPT

ACCEPTED MANUSCRIPT

Links between contaminant hotspots in low flow estuarine systems and altered sediment biogeochemical processes.

3 4

Michael Sutherland1,3, Katherine A. Dafforn1,2*, Peter Scanes3, Jaimie Potts3, Stuart L. Simpson4, Vivian X.-Y. Sim1,2 and Emma L. Johnston1,2 1 Applied Marine and Estuarine Ecology Laboratory, School of Biological, Earth and Environmental Sciences, University of New South Wales, Sydney, NSW 2052, Australia 2 Sydney Institute of Marine Science, Mosman, NSW 2088, Australia 3 Coastal Waters Unit, Science Division, New South Wales Office of Environment and Heritage, Sydney, NSW 2022, Australia 4 Centre for Environmental Contaminants Research, CSIRO Land and Water, Sydney, NSW 2232, Australia.

SC

M AN U

Key words: - Stormwater - Retention – Urbanisation – Biogeochemical Cycling – Benthic Flux – Nutrient Cycling

TE D

Abstract: The urbanisation of coastal zones is a major threat to the health of global estuaries and has been linked to increased contamination (e.g. metals) and excess organic matter. Urban stormwater networks collect and funnel contaminants into waterways at point sources (e.g. stormdrains). Under dry, low flow conditions, these stormwater contaminants can accumulate in sediments over time and result in modifications to benthic sediment biogeochemical processes. To quantify these processes, this field study measured differences in benthic metabolism (CR, GPP, NEM) and sediment-water nutrient fluxes (NH3, NOx, PO4) associated with stormdrains (0 m, 200 m and 1000 m away) and increased waterretention (embayments vs channels). Significant changes to benthic metabolism were detected with distance from stormdrains, and with differences in water-retention rates, above natural spatial and temporal variation. Oxygen consumption was ~50% higher at stormdrains (0 m) compared to 1000 m away and >70% higher at stormdrains (0 m) located in embayments compared to channels. Oxygen production also appeared to decrease with distance from stormdrains in embayments, but patterns were variable. These changes to benthic metabolism were of a magnitude expected to influence benthic nutrient cycling, but NH3, NOx and PO4 fluxes were generally low, and highly spatially and temporally variable. Overall, metal (Cu) contamination explained most of the variation in sediment biogeochemical processes between embayments and channels, while sediment grain size explained differences in fluxes with distance from stormdrains. Importantly, although there was evidence of increased productivity associated with stormdrains, we also detected evidence of early hypoxia suggesting that systems with legacy stormwater contaminants exist on a tipping point. Future work should investigate changes to sediment processes after a major rainfall event, when large and sudden inputs of potentially toxic contaminants occur. Monitoring benthic O2 fluxes could be a sensitive measure of ecological change under these conditions.

EP

19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43

*Corresponding author: Katherine Dafforn, School of Biological, Earth and Environmental Sciences, University of New South Wales, Sydney, NSW 2052, Australia, +61 2 93858701, [email protected].

AC C

5 6 7 8 9 10 11 12 13 14 15 16 17 18

RI PT

1 2

44 45

1

ACCEPTED MANUSCRIPT

1. Introduction

47 48 49 50 51 52 53 54 55 56 57 58 59 60 61 62

Estuaries are under increasing pressure from human activities and as a result are among the most heavily degraded coastal ecosystems (Lotze et al. 2006, Cloern et al. 2016). Many estuarine habitats have been highly modified by encroaching coastal infrastructure and the development of catchments for urban, industrial, and agricultural uses (Newton et al. 2012, Sekovski et al. 2012). The stormwater infrastructure (i.e. drains/channels/culverts) servicing these catchments introduce a multitude of anthropogenic wastes to aquatic systems, transported in surface-runoff over an increasingly impervious environment (Beck and Birch 2014, Coates-Marnane et al. 2016, Sharley et al. 2016). This runoff contains a complex mix of materials (e.g. sediments, organic matter (OM), inorganic nutrients, metal(loid)s and organic chemicals) many of which are considered contaminants to receiving estuarine basins (Kennish 2002, Laursen et al. 2002). Where these contaminants accumulate at high concentrations in estuarine sediments, they are capable of interfering with important ecosystem functions (e.g. fluxes of energy or material through productivity, decomposition and nutrient cycling (Srivastava and Vellend 2005)) through toxic or enriching effects (Long et al. 1995, Worm et al. 2006, Banks et al. 2013, Chapman et al. 2013, von Glasow et al. 2013).

63 64 65 66 67 68 69 70 71 72 73 74 75 76 77

Maintaining sediment health is particularly important in shallow estuaries where well-lit sediments allow benthic biogeochemical processes to dominate ecosystem functions like primary productivity, carbon cycling, and the remineralisation of macronutrients (primarily nitrogen, phosphorus) (Ferguson and Eyre 2013). Toxic contaminants (e.g. metals) introduced from point sources such as stormdrains accumulate in the soft-sediments of these estuaries (Chariton et al. 2010, Dafforn et al. 2012) as a result of their characteristically slow-flowing water regimes, potentially increasing the exposure potential for benthic communities. Natural organic matter carried in stormwater and deposited in slow-flowing environments can trigger a progression of eutrophic symptoms in these estuaries, shifting system productivity from benthic to pelagic-dominated (Bowen and Valiela 2001) with consequences for submerged vegetation, lowered dissolved oxygen concentrations and the proliferation of potentially harmful algal species (Catford et al. 2007, Bricker et al. 2008). The build-up of legacy contaminants from stormwater inputs creates a multiple stressor scenario, where sediment communities are exposed to a range of contaminants that may result in significant ecological change (Dafforn et al. 2016).

78 79 80 81 82 83 84 85 86 87 88 89

Sediment communities, including benthic microbes, algae and infauna, are important drivers of ecosystem function in estuarine habitats (Kristensen and Blackburn 1987, MacIntyre et al. 1996, Savage et al. 2012, Pratt et al. 2014). These communities influence total ecosystem metabolism, with benthic microalgae (BMA) accounting for between 5% to >50% of overall primary productivity (Raalte et al. 1976, Sundbäck and Jönsson 1988) and ~25% of the organic matter respired in these systems (Nixon 1981). Interactions between micro- and macro-sediment organisms can also influence ecosystem metabolism with benthic infauna acting to enhance the availability of particulate organic matter to BMA through grazing and defecation, as well as the oxygenation and destablisation of surficial sediments through burrowing (Kristensen and Blackburn 1987, Laursen et al. 2002, Banks et al. 2013, Harris et al. 2016). The decomposition (mineralisation) of organic matter by benthic heterotrophs also produces a number of important nitrogen compounds, which

AC C

EP

TE D

M AN U

SC

RI PT

46

2

ACCEPTED MANUSCRIPT

along with carbon and phosphorus are essential to biomass production by autotrophs (Ferguson et al. 2003, Howarth and Marino 2006). Nitrification reduces sediment acidification due to ammonia/ammonium accumulation, but critically, the products (collectively known as NOx) drive removal of excess nitrogen from estuaries by microbial denitrification (Seitzinger 1988). Denitrification essentially removes nitrogen from sediments by reducing NOx into relatively inert dinitrogen gas (N2). The diversity and abundance of, and interactions between, sediment dwelling organisms, including microbes, are therefore significant drivers of estuarine nutrient cycles (Maher and Eyre 2010, Eyre et al. 2011) and changes associated with legacy stormwater contaminants could have severe ecological consequences.

100 101 102 103 104 105 106 107 108 109 110 111 112 113 114 115 116 117 118 119 120 121 122 123 124 125 126 127 128

Higher rates of organic matter deposition can increase benthic community respiration rates, depleting oxygen (O2) in both the benthos and overlying water column (Kristensen and Blackburn 1987, Eyre and Ferguson 2002). When O2 is low, nitrification is limited and this subsequently limits the supply of oxidised nitrogen (NOx) to denitrifiers (Kristensen and Blackburn 1987). The resulting build-up of ammonia in sediment pore water can then flux via diffusion into the water column and become available to stimulate growth of pelagic microalgae (Laursen et al. 2002). This enhanced coupling between benthic-pelagic processes promotes a series of positive feedback mechanisms, which further contribute to the deterioration of proper biogeochemical functioning (Krom and Berner 1980). Elevated concentrations of ammonia in pore waters may also result in direct toxicity to some species and contribute to broader impacts to benthic communities (Batley and Simpson 2009). Toxic effects of metal contamination include community shifts from loss of sensitive species and increased abundances of tolerant species (Dafforn et al. 2013). Where these are functionally important, this can translate to changes in biogeochemical process rates such as metabolism, denitrification and nitrogen regeneration (Banks et al. 2013).This study investigated biogeochemical processes in sediment communities at locations in Port Jackson, Australia, with a legacy of stormwater contamination. Locations were either poorly flushed (high retention embayments) or well flushed (low retention open channels) to investigate whether retention has the potential to influence sediment processes. Within each location, sampling was undertaken at sites close to large stormdrains, which were expected to have the highest levels of contaminants and organic matter from a history of heavy sedimentation events, and at sites at increasing distances from the outlets. Measurements of benthic metabolism (community respiration, gross primary productivity and net ecosystem metabolism) and sediment-water nutrient fluxes were compared between channels and embayments and with increasing distance from storm drains. We expected to find multiple stressor hotspots at the sites closest to stormdrains and within poorly flushed embayments, and that these would negatively influence sediment biogeochemical processes compared to reference sites 1000 m away from the point source or in well-flushed channels.

129

2. Materials and Methods

130

2.1. Study Area

131 132 133 134 135 136

Port Jackson (Sydney Harbour) is a large temperate estuary situated on the south-east coast of New South Wales, Australia (33o50‘S, 151o15‘E) (Figure 1). The estuary is a drowned river valley (Roy et al. 2001) and comprises a complex network of inlets and embayments connected through channels adjacent to a winding and increasingly larger channel opening to the Tasman Sea. Activities such as land reclamation and the dredging of shipping channels have extensively modified the Harbour’s bathymetry and shoreline over the past

AC C

EP

TE D

M AN U

SC

RI PT

90 91 92 93 94 95 96 97 98 99

3

ACCEPTED MANUSCRIPT

200 years; 110 km2 (22%) of the estuary has been filled in and 60 km (20%) of shoreline lost since the late 1800’s (Birch and McCready 2009). Over 90% of the Sydney Harbour catchment is either urban, or industrial (Taylor et al. 2004). Historical industrial practices and past urban stormwater inputs have left a legacy of contamination in surrounding sediments (Chariton et al. 2010, Dafforn et al. 2012). Copper, lead and zinc are 10×, 20×, and 90× that of pre-anthropogenic concentrations respectively (Birch and Taylor 1999).

143 144 145 146 147 148 149 150 151

Four locations in the central portion of Sydney Harbour were selected for study: Hen and Chicken Bay, Iron Cove, Lane Cove River, and the Upper Parramatta River (Figure 1). These locations were selected based upon a number of criteria: (i) depths averaging between 1 – 1.5 m, (ii) similar sediment grain sizes, and (iii) similar temperature and salinity due to their central position in the estuary (Dafforn et al. 2012). Locations were separated by water retention characteristics (channels and embayments). The Parramatta and Lane Cove River locations are well flushed channels with low retention and high flow. Iron Cove and Hen and Chicken Bay experience very little flushing and were therefore selected as high-retention sites (Lee et al. 2011).

152

2.2. Sample Design and Collection

153 154 155 156 157 158 159 160 161 162 163 164

Sediment samples were collected once a month during the Australian summer between February and May 2014 from three sites in each of the four locations (described above). Sites were 0, 200 or 1000 m distance from a major stormdrain. These distances were based upon hydrodynamic modelling by Lee et al. (2011) who described freshwater-plume penetration into the Port Jackson estuary during significant rainfall events (>50 mm/day). We expected legacy contaminant concentrations to be highest at sites exposed in the past to major stormwater inputs and reduced with distance away. Sampling was done during dry conditions (<5 mm rainfall/day) and therefore ongoing stormwater inputs into the estuary were minimal (<0.1 m3/s) (Birch and Rochford 2010) and unlikely to have influenced ecological or environmental patterns. Sampling was done 4 times to investigate if patterns related to retention and distance were consistent over time and detectable above natural temporal variation.

165 166 167 168 169 170 171 172

Polycarbonate cores (8 cm inner diam. × 40 cm length) were used to collect sediment from a depth of ~1.5 m at each site. Cores were 30 cm in height with approximately 0.5 L of overlying water. Four replicate cores were collected from each site (n = 12 per location), with care taken to reposition the corer with each collection, and to discard any sample which showed previous surface disturbance. In addition, two replicate sediment samples for analyses of metals, carbon and nitrogen, and grain size were collected from the same sites using a Van Veen grab. Surficial sediments from cores or grabs were homogenised and subsampled then stored on ice before transfer to a -20 oC freezer until processing.

173 174 175

Rainfall data were accessed from the Bureau of Meteorology using Observatory Hill as a representative station for Sydney Harbour. Temperature and salinity were measured in situ with a YSI-Sonde 6600 during each sampling period at a depth of 0.5 m.

176

2.3. Biogeochemical Processes

177 178 179 180 181

Benthic metabolism and sediment/water nutrient fluxes were measured with sediment core incubations under laboratory conditions (Eyre and Ferguson 2002, Ferguson et al. 2003). Incubations of collected sediment cores (48/sampling month) were done over a 48 h period using two closed-system incubation chambers each able to hold 12 sediment cores. As a means of isolating light and non-light dependant reactions, each core was incubated

AC C

EP

TE D

M AN U

SC

RI PT

137 138 139 140 141 142

4

ACCEPTED MANUSCRIPT

sequentially under simulated-light conditions and then in darkness, as per the NICE protocol (Dalsgaard et al. 2000). During these incubations external variables like temperature and light were controlled to reflect conditions measured in situ during sediment sampling. Sediment cores were sealed and submerged in a radial arrangement beneath temperaturecontrolled freshwater. During light incubations, cores were illuminated by photosynthetically active radiation (PAR) at intensities similar to those measured in situ (~180 –250 µmol m−2 s−1) delivered by a 1000 W high-pressure sodium bulb suspended directly above the centre of each chamber. Dark incubations were run in total darkness by completely covering each chamber.

191 192 193 194 195 196 197 198 199 200 201 202 203 204 205 206 207 208 209 210

Prior to incubation, cores were sealed for one hour with an acrylic lid and left submerged in each chamber. During this period external water was warmed or cooled via a thermostatically-controlled pump to reflect temperatures measured in-situ and the internal overlying water of each core was gently flushed by a gravity-fed source of estuarine water collected from each respective location. This allowed sediment cores to re-equilibrate to natural temperature and light conditions and to ensure uniform conditions within replicate cores prior to the measurement of benthic processes. Throughout both acclimation periods and the following incubations, internal overlying water in the core was continuously stirred using magnetised strips suspended vertically from each core’s acrylic lid to avoid stratification. These magnetic strips interact with the magnetised ends of a rotating bar at the centre of each chamber, causing a ‘flicking’ motion every few seconds. Each light and dark incubation was run over 2.5 h with samples and measurements taken at the start and end of the incubation period. This time period was determined on the basis of previous studies (Ferguson et al. 2003) to be sufficient for measurable changes to water chemistry, without triggering longer-term artefacts due to the over-consumption of O2 to levels below 20 % of initial concentrations (Dalsgaard et al. 2000). Dissolved oxygen concentrations (±0.01 mg L-1) and internal water temperature were measured directly via a port in each core’s lid using a Hach Optical oxygen/temperature probe. Gross benthic primary productivity was calculated by subtracting dark O2 flux rates (respiration) from net light flux rates (respiration + productivity).

211 212 213 214 215 216 217 218

Thirty millilitre water samples were also collected at each sample time which were then filtered through 45 µm cellulose acetate filters into clean polypropylene tubes and immediately stored at -20 oC for later nutrient analysis. Nutrient concentrations were measured by technical staff at the Lidcombe laboratories of the New South Wales Government’s Office of Environment and Heritage using flow injection analysis (LachatTM QuikChem 8000). Ammonia (NH3) was determined by the automated phenate method, phosphorus (PO4) by the automated ascorbic acid reduction method and NOx (nitrite + nitrate) by automated cadmium reduction.

219 220

The benthic flux of nutrients across the sediment/water interface were calculated using the following formula:

AC C

EP

TE D

M AN U

SC

RI PT

182 183 184 185 186 187 188 189 190

221 222

BF = ([Ct1 – Ct0] x V/SA)/T

223 224 225

where BF = benthic flux of nutrients (mmol m-2 h-1), Ct0 = nutrient concentration at incubation start (mmol L-1), Ct1 = nutrient conc. at incubation end (mmol L-1), V = volume of

5

ACCEPTED MANUSCRIPT

overlying water in the core (L), SA = surface area of sediment in the core (m2), and T = time of incubation (h).

228 229 230 231 232

Each incubation included two blank cores per location. These cores contained no sediment, but were otherwise treated identically to sediment-containing cores. The blank cores were included to estimate the potential for metabolism and nutrient cycling driven primarily by phytoplankton in the water column. These measurements were then subtracted from total O2/nutrient fluxes in order to isolate benthic flux rates.

233

2.4 Sediment Characteristics

234 235 236 237 238 239 240 241 242 243

Briefly, metals analyses followed Dafforn et al. (2012) with sediments oven-dried (50 oC) and homogenised to a fine powder with mortar and pestle before microwave digestion according to Method 3051A (USEPA 2007). Following digestion metal concentrations were analysed using ICP-AES (Perkin Elmer, Optima7300DV, USA). Grain size analyses were done on wet sediments using a Malvern laser particle size analyser (Mastersizer 2000) and the percentage <63 µm calculated for model inputs. Total organic carbon and total nitrogen were analysed at Isoenvironmental (South Africa) using 5-20 mg of dried, homogenised sample in a 20-20 IRMS linked to an ANCA SL elemental analyser (Europa Scientific). Details of the mean and range for each variable are included in supplementary materials (Supplementary Table S1).

244

2.5. Statistical Analysis

245 246 247 248 249 250 251 252 253 254 255 256 257 258 259 260 261 262 263 264 265

Patterns in sediment biogeochemical processes (metabolism and nutrient cycling) in relation to retention and distance were analysed using permutational analysis of variance (permANOVA). Four factors were included in the design; Sampling month (random factor with four levels: February, March, May and June), Retention type (fixed factor with two levels: channels and embayments), Location (random factor with four levels nested in Retention type: Hen and Chicken Bay, Iron Cove, Upper Parramatta River and Lane Cove River), Distance from stormdrain (fixed factor with three levels: 0, 200 and 1000 m). Measures of community respiration, gross primary productivity, net primary productivity/ecosystem metabolism, and nutrient fluxes (NH3, NOx and PO4) were analysed with this design. Pairwise comparisons were done for significant main effects or significant interaction terms. Full details of comparisons are provided in supplementary materials (Supplementary Table S2) and summary p-values included in text. Fluxes were also modelled against sediment characteristics with forward selection distance-based linear modelling (DistLM) and visualised with distance-based redundancy analysis (dbRDA) (Anderson 2001, McArdle and Anderson 2001). Multivariate datasets were initially explored with a Draftsman Plot and where correlations were > 0.9 then a single variable was selected as representative for further analysis (Al was also representative of As, Fe, Ni and Pb was also representative of Zn). All data were normalised to account for differences in the order of magnitude of chemical variables and analyses were done on resemblance matrices constructed using Euclidean distance. Analyses were done using the PRIMER 6 statistical package with the PERMANOVA+ add-on (PRIMER-E, Plymouth Marine Laboratory, UK).

266

3. Results

267

3.1 Physico-chemical Conditions

268 269

Rainfall in Sydney Harbour was < 100 mm across all months of the sampling period (Feb – May, 2014) (Bureau of Meteorology, 2014). Salinities remained within estuarine ranges

AC C

EP

TE D

M AN U

SC

RI PT

226 227

6

ACCEPTED MANUSCRIPT

throughout the survey (31 – 36). Water temperatures were typical of a temperate estuary, ranging from 25 oC in February to 17 oC in May.

272

3.2. Benthic Metabolism

273 274 275 276 277 278 279 280 281

Benthic oxygen (O2) fluxes during dark incubations were used as a measure of overall benthic community respiration (CR). Community respiration patterns varied with water retention characteristics of the location and increasing distance from stormdrains (Re x Di, p < 0.05; Table 1). Specifically, community respiration was ~50 % greater at 0 m distance than 1000 m distance for both embayments and channels (pairwise comparisons, p < 0.05, Figure 2a). Furthermore, community respiration was >70 % greater in embayments than channels (Figure 2a), but only if sites were close to stormdrains (0 m and 200 m) (pairwise comparisons, p < 0.05, Figure 2a). Oxygen consumption was also spatially and temporally variable (Ti × Lo(Re), p < 0.05; Table 1).

282 283 284 285 286 287 288 289 290 291 292

Gross primary productivity (GPP) was determined by subtracting dark (CR) from light O2 flux rates, thereby giving the total potential for O2 production during light incubations. Benthic GPP differed significantly between retention types (Re, p < 0.05; Table 1), with higher rates of GPP at sites within embayments than channels (Figure 2b). Gross primary productivity also differed with distance from the stormdrain and time, although both showed variability among locations (Ti × Lo(Re) × Di, p < 0.05; Table 1). GPP was highest at the site closest to the stormdrain for all sampling months in Iron Cove (pairwise comparisons, p < 0.05, Figure 2b). Patterns of decreasing GPP away from storm drains were also observed in Hen and Chicken Bay, although the difference was only significant at one sampling time (pairwise comparisons, p < 0.05, Figure 2b). GPP in channel locations was more variable with patterns rarely significant and not consistent between months (Figure 2b).

293 294 295 296 297 298 299 300 301 302

Rates of O2 flux measured during light incubations were used as a measure of net primary productivity (NPP = GPP - CR). Benthic NPP differed significantly between retention types (Re, p < 0.05; Table 1) and higher rates of O2 surplus (efflux) were observed in embayments than in channels (Figure 2c). NPP also differed with distance from the stormdrains in embayments, but this was variable between locations and sampling times (Ti × Lo(Re) × Di, p < 0.05; Table 1). Specifically, during February, March, and April, O2 production was lowest at sites 0 m from stormwater drains, and highest at sites 200 m from the drains in Hen and Chicken Bay (pairwise comparisons, p < 0.05, Figure 2c). Iron Cove displayed similarly high values for sites at 200 m during April, but experienced highest NPP at sites at 0 m during February, March, and May (pairwise comparisons, p < 0.05, Figure 2c).

303 304 305 306 307 308 309 310 311 312

The average daily O2 flux (CR and NPP each multiplied by 12hrs) was used to infer net ecosystem metabolism (NEM). Patterns of daily oxygen transfer between the sedimentwater were both spatially and temporally variable (Ti × Lo(Re) × Di, p < 0.05; Table 1). Low retention locations (channels) acted as net sinks for O2 across all months (aside from some ecologically negligible outputs during March), excluding May when Upper Parramatta River was a net source of O2 (Figure 2d). High retention locations (embayments) displayed much greater temporal and spatial variation in net O2 fluxes, tending towards sinks in February and March, and sources in April and May (Figure 2d). Sediments in Hen and Chicken Bay (high retention) showed the highest rates of O2 uptake, which were greater at sites closest to stormdrains than those furthest away during February and May (Figure 2d).

313

3.3. Benthic Nutrient Cycling

AC C

EP

TE D

M AN U

SC

RI PT

270 271

7

ACCEPTED MANUSCRIPT

Dark benthic NH3 fluxes were very low, but almost always positive (i.e. out of the sediments, Figure 3a) and varied with distance and between retention types, but patterns were only significant during some sampling months (Ti x Re and Ti x Di, p < 0.05; Table 2). Although pair-wise comparisons were unable to detect consistent significant differences in spatial and temporal effects on NH3 fluxes between sediment and water (pairwise comparisons, p > 0.05), NH3 release from sediments under dark conditions appeared to decrease with distance for all locations in February (Figure 3a), and all locations excluding Hen and Chicken Bay in April (Figure 3a). Rates of benthic NH3 fluxes under light conditions were spatially and temporally variable, but showed no significant differences as a function of retention type or with distance from stormwater drains (Ti x Lo(Re), p < 0.05; Table 2). In general, rates of NH3 flux were low across all locations and sampling periods.

325 326 327 328 329 330 331 332

Dark and light benthic nitrite + nitrate (NOx) fluxes were generally low and negative in direction in all locations over time (Figure 3c). Significant spatial and temporal variation was observed in NOx fluxes under both lighting conditions (Ti x Lo(Re), p < 0.05; Table 2). Dark NOx flux rates also varied significantly with retention type, but only at certain distances from stormwater source (Re x Di, p < 0.05; Table 2). Pair-wise comparisons failed to detect any significant retention by distance differences in dark NOx flux (pairwise comparisons, p > 0.05, Figure 3c), however there was some evidence of decreasing NOx flux rates away from stormwater in channels (Figure 3c).

333 334 335 336 337 338 339

Rates of benthic phosphate (PO4) fluxes (dark and light) varied significantly with distance from stormwater source (Di, p < 0.05; Table 2, Figure 3e, f). However, differences in PO4 flux rates were near indistinguishable given their small magnitude (<±5 µmol-2 h-1) across most locations for a majority of temporal replicates (Figure 3e,f). Phosphate fluxes also differed among times (Ti, p < 0.05; Dark) and among times and locations (Ti x Lo(Re), p < 0.05; Light) (Table 2) and dark PO4 flux rates appeared highest across all locations in February (Figure 3e).

340

3.4. Variation in Benthic Processes related to Sediment Contamination and Characteristics

341 342 343 344 345 346 347 348 349 350 351

Differences in flux measurements were fitted to metal concentrations, total organic carbon, total nitrogen and grain sizes (% < 63 µm) analysed from associated sediments (Table 3, Figure 4a, b). Channels and embayments separated out along two apparent axes and within embayments there was further separation between 0 m and 200 m from the stormdrains and 1000 m away (Figure 4a). Copper concentrations and sediment grain size explained the most variation in the fluxes (Table 3) with Cu particularly good at discriminating between locations with Low and High retention and between sites 0 m/200 m and 1000 m (explaining 12 % of the variance) from stormdrains. Grain size was a better discriminator among sites (0, 200 and 1000 m) within channels (Figure 4b). In total, the measured variables explained 23 % of the variation in the flux data, but only Cu, Cr and grain size contributed significantly to the fitted model (Table 3).

SC

M AN U

TE D

EP

AC C

352

RI PT

314 315 316 317 318 319 320 321 322 323 324

353

4. Discussion

354 355 356 357 358

Urban stormwater contamination in estuaries has the potential to significantly alter benthic biogeochemical processes such as metabolism and nutrient cycling. We expected that the accumulation of organic matter enrichment and metal contaminants in our poorly flushed systems would act negatively on natural processes, but instead found that community respiration was significantly higher adjacent to stormdrains and patterns of increased gross

8

ACCEPTED MANUSCRIPT

primary productivity. Together these results suggest enriching effects from legacy stormwater contaminants on benthic heterotrophic communities. However, given the high levels of sediment contamination, often far above sediment quality guideline values (Simpson et al. 2013), this may be an indication that the system is close to eutrophication and we found some evidence of early-stage hypoxia (Howarth et al. 2011). Ongoing monitoring of sensitive indicators such as benthic metabolism, particularly under dry and wet conditions would provide better understanding the functional consequences if additional stress is placed on these shallow estuarine systems. Furthermore, responses were more pronounced in embayments where contaminants are more likely to have been retained over time, compared to channels that experience vigorous tidal and fluvial flushing. The greater magnitude of biological response to anthropogenic contamination in systems with high levels of water retention (slow exchange) suggests greater vulnerability in these environments, and the potential for important processes such as denitrification to be inhibited. These hotspots of contamination in low flow areas should be a priority for ongoing management and remediation efforts.

374 375 376 377 378 379 380 381 382 383 384 385 386 387 388 389 390 391 392 393

The supply of organic matter to sediments is a major factor influencing benthic community respiration in estuaries and other shallow coastal systems (Kristensen and Blackburn 1987, Moran and Hodson 1990). Enhanced benthic respiration at sites closest to stormdrains was likely caused by localised enrichment of organic matter from past stormwater discharges. Greater responses in benthic communities to organic matter exposure in high retention environments were likely to be associated with prolonged residencies of deposited organic materials, hence greatest differences occurring closest to stormdrains. Previous studies have identified similar positive relationships between rates of benthic organic matter decomposition and retention time of organic inputs to estuarine sediments (Hopkinson and Vallino 1995, Paerl et al. 1999). Hydrodynamics plays an important role in determining residency time by controlling the horizontal transport of suspended and resuspended materials through estuarine systems (Vanderborght et al. 1977, Christiansen et al. 1997). Unlike open channels, closed embayments do not experience significant bi-directional water movement driven by fluvial and tidal forces (Asmus et al. 1998). Instead, the absence of directional currents within embayments leads to a radial distribution of contaminants around a stormdrain as we observed in this study at 0 m and 200 m distances. We also observed increases in TOC and TN close to stormdrains in embayments and this accumulation of organic matter may be explained by water current-driven sediment resuspension, which tends to be reduced in sheltered environments (Peckol and Rivers 1996, Martins et al. 2001).

394 395 396 397 398 399 400 401 402 403 404 405

Benthic autotrophs (BMA) benefit from increased organic matter decomposition by being first to receive the mineralised inorganic compounds released during heterotrophic metabolism. It is common, therefore, to find direct correlations between rates of benthic respiration and benthic productivity in eutrophic estuaries (Eyre and Ferguson 2002, Laursen et al. 2002, Ferguson et al. 2003). Coupling between benthic processes was only consistently evident at one of the locations surveyed in this study. The high-retention embayment Iron Cove exhibited the highest rates of GPP closest to the main stormdrain during all months excluding April (where GPP at 0 m = 200 m >1000 m). In contrast, rates of GPP at Hen and Chicken Bay (embayment) were lowest at sites close to stormdrains during February and May, with evidence of coupled heterotrophic-autotrophic enrichment only apparent at sites at 200 m distances from stormdrains during April and May. This indicates that benthic productivity was unlikely to be benefiting from increased organic matter

AC C

EP

TE D

M AN U

SC

RI PT

359 360 361 362 363 364 365 366 367 368 369 370 371 372 373

9

ACCEPTED MANUSCRIPT

mineralisation and this may be due to greater toxic effects from the presence of metal contaminants. Our measures of organic matter enrichment explained less of the variation in sediment biogeochemical processes (2 %) compared to sediment metal concentrations (Cu and Cr) and sediment grain size (17 %). In previous studies, discrepancies between rates of organic matter decomposition and benthic biomass production have been attributed to secondary effects of sediment deoxygenation, primarily in environments identified as being ‘net heterotrophic’ (Eyre and Ferguson 2002).

413 414 415 416 417 418 419 420 421 422 423 424 425 426

When the consumption of O2 exceeds its production by benthic autotrophs, estuarine sediments become net sinks for available water column O2 and are described as being net heterotrophic (CR>NPP). Low-retention locations were found to be net heterotrophic across all locations and months, excluding May where there was evidence of net autotrophy (NPP>CR) in the Upper Parramatta River. This is likely related to the low rates of GPP observed in low-retention channels, which is not uncommon for benthic communities exposed to high hydrodynamic activity. Wave energy is known to be a significant factor affecting BMA biomass production in estuarine habitats (MacIntyre et al. 1996, Cook et al. 2004). BMA exist in the upper few centimetres of sediment, and therefore may be disturbed by turbulence and sheer stresses generated by tidal currents and wind-waves (MacIntyre et al. 1996). Although possible, it is unlikely that instances of net heterotrophy in highretention environments such as those observed at Hen and Chicken Bay were a result of hydrodynamic stress to producers, but were more likely explained by modifications to sediment nutrient cycling.

427 428 429 430 431 432 433 434 435 436 437 438 439 440 441 442 443 444 445 446

In severe cases where complete sediment anoxia occurs, drastic changes to benthic nutrient cycling can result, primarily in oxygen-dependant processes such as nitrification (Figure 4) (Kristensen and Blackburn 1987). This requires oxygen depletion of overlying waters. Net heterotrophic sediments should release significant amounts of inorganic nitrogen into the overlying water column due to enhanced benthic respiration (ammonification) in the absence of oxidising nitrifiers (Kemp et al. 1990, Laursen et al. 2002). There was little evidence of significant changes to benthic nutrient cycling in response to elevated benthic respiration. Sediment-water NH3 fluxes varied in both magnitude and direction during daylight, however, dark fluxes were seen to differ significantly both between retention types, and with distance from the stormwater source. These changes were only identified during particular months, and for the most part did not correlate with instances of significantly enriched benthic respiration. The exception to this was during February, where net heterotrophy in all locations correlated with increased dark NH3 efflux at sites closest to stormwater outlets. Other than this single event, there were large deficits in the potential amount of organic nitrogen ammonified (assuming that the decomposition of algae followed the ‘Redfield’ ratio of C:N:P as 106:16:1; (Hillebrand and Sommer 1999)) compared to that which effluxed into the water column. This, coupled with the fact that sediments also actively took up NOx throughout the study at sites with elevated benthic respiration, may indicate that significant amounts of nitrogen are being either immobilised in the benthic biomass, or removed via denitrification (Ferguson et al. 2003).

447 448 449 450 451 452

These findings indicate that, although some highly heterotrophic communities exist in proximity to stormwater outlets, BMA are still able to mediate the provision of macronutrients to the pelagic zone through assimilation. This explanation does not, however, explain why enrichment of GPP was not always apparent with distance from stormwater sources, particularly in Hen and Chicken Bay where there appeared to be a negative correlation between GPP and CR. It is possible that denitrification could be driving

AC C

EP

TE D

M AN U

SC

RI PT

406 407 408 409 410 411 412

10

ACCEPTED MANUSCRIPT

nitrogen removal from these systems, however, even when rates of benthic respiration were double those described by Eyre and Ferguson (2002) as necessary to exert pressure upon benthic nitrification/denitrifaction (~1500 µmol O2 m-2 h-1; assuming 1:1 of O2 consumption: CO2 production) in a shallow estuary, net effluxes of NH3 were not observed. Potential explanations of these anomalies include the effects of external factors on these processes. Detrital source richness and mixture has been shown to significantly alter benthic metabolism and the degradation of organic matter (Kelaher et al. 2013). Terrestrial materials are known to contain significantly different lignin contents, and ratios of labile and refractory compounds to those contained within marine macrophytes and microalgae (Hopkinson et al. 1998). Deposited detritus from terrestrial plants is known to contain significantly higher C:N ratios than those of marine species such as phytoplankton (Hopkinson et al. 1998). It has therefore been proposed that benthic oxidation of terrestrial organic matter tends to drive coastal systems towards net heterotrophy, as the remineralised nitrogen that escapes denitrification cannot support equivalent production as organic matter with a low C:N ratio (e.g. BMA) (Hopkinson and Vallino 1995). This could also explain the differences observed in net metabolisms at sites 0 and 200 m from stormwater inputs at Hen and Chicken Bay, because heavy terrestrial detritus would be expected to settle close to discharge outlets, while dissolved inorganic nitrogen is able to be transported further, potentially feeding the production of low C:N autochthonous organic matter. Alternatively, benthic respiration rates may have been over-estimated. A problem with estimating benthic production from changes in O2 fluxes is the extrapolation of dark respiration rates to the light period, because O2 production may modify dark O2 consumption through photorespiration, the release of DOC from phototrophs and enhanced nitrification (Eyre and Ferguson 2002).

477 478 479 480 481 482 483 484 485 486 487 488 489 490 491

Process measurements were highly spatially and temporally variable which may have reduced the ability to detect main effects. Other sediment characteristics and contaminants were not quantified in this study and may have added to the variation that could not be explained by our model (Simpson et al. 2004, Spruill and Bratton 2008). Although significant changes to benthic nutrient cycling were not observed during this study, significant modifications to benthic metabolism did occur in sediments in close proximity to stormdrains and embayments with high water retention. Given that these responses were observed during what would be considered dry conditions, there is significant potential for greater effects should higher rainfalls occur, particularly if they coincided with warmer months (Spring - Summer). Higher rates of organic matter deposition would further fuel benthic respiration and the subsequent release of heterotrophic metabolites into surficial sediments. Increased turbidity as a result of sediment resuspension and the introduction of high amounts of suspended particulate matter could potentially impede the mediatory role exhibited by benthic autotrophs in nutrient releases to the pelagic zone, causing a shift in stable states.

SC

M AN U

TE D

EP

AC C

492

RI PT

453 454 455 456 457 458 459 460 461 462 463 464 465 466 467 468 469 470 471 472 473 474 475 476

493

5. Conclusion

494 495 496 497 498 499

Legacy stormwater contaminants in shallow estuarine waters significantly altered benthic metabolism, and could potentially affect the biogeochemical cycling of nutrients. A clear relationship between retention and enriched benthic heterotrophy was observed in this study, which led to elevated rates of oxygen consumption that exceeded benthic primary production. Despite this, sediment communities within the sampled locations appeared to function relatively well, with no evidence of mass effluxes of growth-promoting

11

ACCEPTED MANUSCRIPT

macronutrients into overlying water bodies, even at sites closest to contaminant sources. There was evidence, however, of early-stage hypoxia in some locations where benthic metabolisms were net heterotrophic, but only during warmer months.

503 504 505 506 507 508 509 510

The observed differences in biological responses to contaminant legacies between retention types could have important applications for future estuarine management strategies. The elevated rates of benthic enrichment observed in high-retention locations suggests that embayments are particularly vulnerable to the degradative effects of contaminant retention, while the apparent resilience of fast-flowing channels may mean that these locations are more suitable for handling anthropogenic wastes. As coastal cities grow, the management of legacy and ongoing contaminant human inputs into systems such as estuaries will be important to maintain the ecosystem functions that estuaries provide.

511 512 513 514 515 516 517 518 519 520 521

Many urbanised estuaries employ long-term physical management strategies to control stormwater discharges from major contaminant source-points (e.g. sediment traps, net filtration) (Madhani and Brown 2015). There have also been some positive moves towards blue-green ecological engineering such as the construction of stormwater wetlands and other forms of biofilters (Greenway in press). Given the high levels of contamination we measured in sediments adjacent to stormdrains and the related changes to sediment functions it does not appear to be enough to only control further inputs. Management efforts to remediate sediments e.g. through dredging, capping or restoration of natural communities (Zarull et al. 2002, Seidel et al. 2004, Mackie et al. 2007, Francingues et al. 2008, Viana et al. 2008) could go some way to addressing the legacy of stormwater contaminants.

522

Acknowledgements

523 524 525 526 527 528

This research was funded by an ARC Linkage Grant (LP130100364) awarded to ELJ, SLS and PS. Fieldwork was conducted under NSW Department of Primary Industries permit number P13/0007-1.0. The authors would like to thank Elisa Tan and Jamie-Louise Morrison for their lab assistance and Simone Birrer, Tim Lachnit, volunteers and members of the Applied Marine and Estuarine Ecology laboratory for their field assistance. We thank Kate Spencer and two anonymous reviewers whose feedback improved the manuscript.

529 530 531 532 533 534 535 536 537 538 539 540 541 542 543 544 545

ReferencesAnderson, M. J. 2001. A new method for non-parametric multivariate analysis of

EP

TE D

M AN U

SC

RI PT

500 501 502

AC C

variance. Austral Ecology 26:32-46. Asmus, R., M. Jensen, K. Jensen, E. Kristensen, H. Asmus, and A. Wille. 1998. The role of water movement and spatial scaling for measurement of dissolved inorganic nitrogen fluxes in intertidal sediments. Estuarine, Coastal and Shelf Science 46:221-232. Banks, J. L., D. J. Ross, M. J. Keough, C. K. Macleod, J. Keane, and B. D. Eyre. 2013. Influence of a burrowing, metal-tolerant polychaete on benthic metabolism, denitrification and nitrogen regeneration in contaminated estuarine sediments. Marine Pollution Bulletin 68:30-37. Batley, G. E. and S. L. Simpson. 2009. Development of guidelines for ammonia in estuarine and marine water systems. Marine Pollution Bulletin 58:1472-1476. Beck, H. and G. Birch. 2014. The nature and source of irregular discharges to stormwater entering Sydney estuary, Australia. Environmental Pollution 188:172-176. Birch, G. and L. Rochford. 2010. Stormwater metal loading to a well-mixed/stratified estuary (Sydney Estuary, Australia) and management implications. Environmental Monitoring and Assessment 169:531-551. Birch, G. and S. Taylor. 1999. Source of heavy metals in sediments of the Port Jackson estuary, Australia. Science of the Total Environment 227:123-138.

12

ACCEPTED MANUSCRIPT

EP

TE D

M AN U

SC

RI PT

Birch, G. F. and S. McCready. 2009. Catchment condition as a major control on the quality of receiving basin sediments (Sydney Harbour, Australia). Science of the Total Environment 407:2820-2835. Bowen, J. L. and I. Valiela. 2001. The ecological effects of urbanization of coastal watersheds: historical increases in nitrogen loads and eutrophication of Waquoit Bay estuaries. Canadian Journal of Fisheries and Aquatic Sciences 58:1489-1500. Bricker, S. B., B. Longstaff, W. Dennison, A. Jones, K. Boicourt, C. Wicks, and J. Woerner. 2008. Effects of nutrient enrichment in the nation's estuaries: A decade of change. Harmful Algae 8:21-32. Catford, J., C. Walsh, and J. Beardall. 2007. Catchment urbanization increases benthic microalgal biomass in streams under controlled light conditions. Aquatic Sciences 69:511-522. Chapman, P. M., F. Wang, and S. S. Caeiro. 2013. Assessing and managing sediment contamination in transitional waters. Environment International 55:71-91. Chariton, A. A., A. C. Roach, S. L. Simpson, and G. E. Batley. 2010. Influence of the choice of physical and chemistry variables on interpreting patterns of sediment contaminants and their relationships with estuarine macrobenthic communities. Marine And Freshwater Research 61:1109-1122. Christiansen, C., F. Gertz, M. Laima, L. C. Lund-Hansen, T. Vang, and C. Jürgensen. 1997. Nutrient (P, N) dynamics in the southwestern Kattegat, Scandinavia: sedimentation and resuspension effects. Environmental Geology 29:66-77. Cloern, J. E., P. C. Abreu, J. Carstensen, L. Chauvaud, R. Elmgren, J. Grall, H. Greening, J. O. R. Johansson, M. Kahru, E. T. Sherwood, J. Xu, and K. Yin. 2016. Human activities and climate variability drive fast-paced change across the world's estuarine–coastal ecosystems. Global Change Biology 22:513-529. Coates-Marnane, J., J. Olley, J. Burton, and A. Grinham. 2016. The impact of a high magnitude flood on metal pollution in a shallow subtropical estuarine embayment. Science of the Total Environment 569:716-731. Cook, P. L., E. C. Butler, and B. D. Eyre. 2004. Carbon and nitrogen cycling on intertidal mudflats of a temperate Australian estuary. I. Benthic metabolism. Marine Ecology Progress Series 280:2538. Dafforn, K. A., E. L. Johnston, A. Ferguson, C. L. Humphrey, W. Monk, S. J. Nichols, S. L. Simpson, M. G. Tulbure, and D. J. Baird. 2016. Big data opportunities and challenges for assessing multiple stressors across scales in aquatic ecosystems. Marine And Freshwater Research 67:393-413. Dafforn, K. A., B. P. Kelaher, S. L. Simpson, M. A. Coleman, P. A. Hutchings, G. F. Clark, N. A. Knott, M. A. Doblin, and E. L. Johnston. 2013. Polychaete richness and abundance enhanced in anthropogenically modified estuaries despite high concentrations of toxic contaminants. PLoS ONE 8:e77018. Dafforn, K. A., S. L. Simpson, B. P. Kelaher, G. F. Clark, V. Komyakova, C. K. C. Wong, and E. L. Johnston. 2012. The challenge of choosing environmental indicators of anthropogenic impacts in estuaries. Environmental Pollution 163:207-217. Dalsgaard, T., L. P. Nielsen, V. Brotas, P. Viaroli, G. J. C. Underwood, D. B. Nedwell, K. Sundbäck, S. Rysgaard, A. Miles, and M. Bartoli. 2000. Protocol handbook for NICE-Nitrogen Cycling in Estuaries: a project under the EU research programme: Marine Science and Technology (MAST III). National Environmental Research Institute, Denmark. Eyre, B., A. P. Ferguson, A. Webb, D. Maher, and J. Oakes. 2011. Denitrification, N-fixation and nitrogen and phosphorus fluxes in different benthic habitats and their contribution to the nitrogen and phosphorus budgets of a shallow oligotrophic sub-tropical coastal system (southern Moreton Bay, Australia). Biogeochemistry 102:111-133. Eyre, B. D. and A. J. P. Ferguson. 2002. Comparison of carbon production and decomposition, benthic nutrient fluxes and denitrification in seagrass, phytoplankton, benthic microalgae- and macroalgae-dominated warm-temperate Australian lagoons. Marine Ecology Progress Series 229:43-59.

AC C

546 547 548 549 550 551 552 553 554 555 556 557 558 559 560 561 562 563 564 565 566 567 568 569 570 571 572 573 574 575 576 577 578 579 580 581 582 583 584 585 586 587 588 589 590 591 592 593 594 595 596 597

13

ACCEPTED MANUSCRIPT

EP

TE D

M AN U

SC

RI PT

Ferguson, A. and B. Eyre. 2013. Interaction of benthic microalgae and macrofauna in the control of benthic metabolism, nutrient fluxes and denitrification in a shallow sub-tropical coastal embayment (western Moreton Bay, Australia). Biogeochemistry 112:423-440. Ferguson, A. J., B. D. Eyre, and J. M. Gay. 2003. Organic matter and benthic metabolism in euphotic sediments along shallow sub-tropical estuaries, northern New South Wales, Australia. Aquatic Microbial Ecology 33:137-154. Francingues, K. E. G., G. A. Burton, R. Norman, Wolfe, Jr., D. R. Danny, J. V. Donna, and R. John. 2008. Evaluating the Effectiveness of Contaminated-Sediment Dredging. Environmental Science & Technology 42:5042-5047. Greenway, M. in press. Stormwater wetlands for the enhancement of environmental ecosystem services: case studies for two retrofit wetlands in Brisbane, Australia. Journal of Cleaner Production. Harris, R. J., C. A. Pilditch, B. L. Greenfield, V. Moon, and I. Kröncke. 2016. The Influence of Benthic Macrofauna on the Erodibility of Intertidal Sediments with Varying mud Content in Three New Zealand Estuaries. Estuaries and Coasts 39:815-828. Hillebrand, H. and U. Sommer. 1999. The nutrient stoichiometry of benthic microalgal growth: Redfield proportions are optimal. Limnology and Oceanography 44:440-446. Hopkinson, C. and J. Vallino. 1995. The relationships among man’s activities in watersheds and estuaries: A model of runoff effects on patterns of estuarine community metabolism. Estuaries 18:598-621. Hopkinson, C. S., I. Buffam, J. Hobbie, J. Vallino, M. Perdue, B. Eversmeyer, F. Prahl, J. Covert, R. Hodson, and M. A. Moran. 1998. Terrestrial inputs of organic matter to coastal ecosystems: An intercomparison of chemical characteristics and bioavailability. Biogeochemistry 43:211234. Howarth, R., F. Chan, D. J. Conley, J. Garnier, S. C. Doney, R. Marino, and G. Billen. 2011. Coupled biogeochemical cycles: eutrophication and hypoxia in temperate estuaries and coastal marine ecosystems. Frontiers in Ecology and the Environment 9:18-26. Howarth, R. W. and R. Marino. 2006. Nitrogen as the limiting nutrient for eutrophication in coastal marine ecosystems: evolving views over three decades. Limnology and Oceanography 51:364-376. Kelaher, B. P., M. J. Bishop, J. Potts, P. Scanes, and G. Skilbeck. 2013. Detrital diversity influences estuarine ecosystem performance. Global Change Biology 19:1909-1918. Kemp, W., P. Sampou, J. Caffrey, M. Mayer, K. Henriksen, and W. R. Boynton. 1990. Ammonium recycling versus denitrification in Chesapeake Bay sediments. Limnology and Oceanography 35:1545-1563. Kennish, M. J. 2002. Environmental threats and environmental future of estuaries. Environmental Conservation 29:78-107. Kristensen, E. and T. Blackburn. 1987. The fate of organic carbon and nitrogen in experimental marine sediment systems: influence of bioturbation and anoxia. Journal of Marine Research 45:231-257. Krom, M. D. and R. A. Berner. 1980. Adsorption of phosphate in anoxic marine sediments1. Limnology and Oceanography 25:797-806. Laursen, A. E., S. P. Seitzinger, R. Dekorsey, J. G. Sanders, D. L. Breitburg, and R. W. Osman. 2002. Multiple stressors in an estuarine system: effects of nutrients, trace elements, and trophic complexity on benthic photosynthesis and respiration. Estuaries 25:57-69. Lee, S. B., G. F. Birch, and C. J. Lemckert. 2011. Field and modelling investigations of fresh-water plume behaviour in response to infrequent high-precipitation events, Sydney Estuary, Australia. Estuarine, Coastal and Shelf Science 92:389-402. Long, E. R., D. D. MacDonald, S. L. Smith, and F. D. Calder. 1995. Incidence of adverse biological effects within ranges of chemical concentrations in marine and estuarine sediments. Environmental Management 19:81-97.

AC C

598 599 600 601 602 603 604 605 606 607 608 609 610 611 612 613 614 615 616 617 618 619 620 621 622 623 624 625 626 627 628 629 630 631 632 633 634 635 636 637 638 639 640 641 642 643 644 645 646 647 648

14

ACCEPTED MANUSCRIPT

EP

TE D

M AN U

SC

RI PT

Lotze, H. K., H. S. Lenihan, B. J. Bourque, R. H. Bradbury, R. G. Cooke, M. C. Kay, S. M. Kidwell, M. X. Kirby, C. H. Peterson, and J. B. C. Jackson. 2006. Depletion, degradation, and recovery potential of estuaries and coastal seas. Science 312:1806-1809. MacIntyre, H., R. Geider, and D. Miller. 1996. Microphytobenthos: The ecological role of the “secret garden” of unvegetated, shallow-water marine habitats. I. Distribution, abundance and primary production. Estuaries and Coasts 19:186-201. Mackie, J. A., S. M. Natali, J. S. Levinton, and S. A. Sañudo-Wilhelmy. 2007. Declining metal levels at Foundry Cove (Hudson River, New York): Response to localized dredging of contaminated sediments. Environmental Pollution 149:141-148. Madhani, J. T. and R. J. Brown. 2015. The capture and retention evaluation of a stormwater gross pollutant trap design. Ecological Engineering 74:56-59. Maher, D. T. and B. D. Eyre. 2010. Benthic fluxes of dissolved organic carbon in three temperate Australian estuaries: Implications for global estimates of benthic DOC fluxes. Journal of Geophysical Research: Biogeosciences (2005–2012) 115. Martins, I., M. Pardal, A. Lillebø, M. Flindt, and J. Marques. 2001. Hydrodynamics as a major factor controlling the occurrence of green macroalgal blooms in a eutrophic estuary: a case study on the influence of precipitation and river management. Estuarine, Coastal and Shelf Science 52:165-177. McArdle, B. H. and M. J. Anderson. 2001. Fitting multivariate models to community data: A comment on distance-based redundancy analysis. Ecology 82:290-297. Moran, M. A. and R. E. Hodson. 1990. Bacterial production on humic and nonhumic components of dissolved organic carbon. Limnology and Oceanography 35:1744-1756. Newton, A., T. J. B. Carruthers, and J. Icely. 2012. The coastal syndromes and hotspots on the coast. Estuarine, Coastal and Shelf Science 96:39-47. Nixon, S. W. 1981. Remineralization and nutrient cycling in coastal marine ecosystems. Pages 111138 in B. J. Neilson and L. E. Cronin, editors. Nutrients and EStuaries. Springer. Paerl, H. W., J. D. Willey, M. Go, B. L. Peierls, J. L. Pinckney, and M. L. Fogel. 1999. Rainfall stimulation of primary production in western Atlantic Ocean waters: Roles of different nitrogen sources and co-limiting nutrients. Marine Ecology Progress Series 176:205. Peckol, P. and J. Rivers. 1996. Contribution by macroalgal mats to primary production of a shallow embayment under high and low nitrogen-loading rates. Estuarine, Coastal and Shelf Science 43:311-325. Pratt, D. R., A. M. Lohrer, C. A. Pilditch, and S. F. Thrush. 2014. Changes in Ecosystem Function Across Sedimentary Gradients in Estuaries. Ecosystems 17:182-194. Raalte, C. D. V., I. Valiela, and J. M. Teal. 1976. Production of epibenthic salt marsh algae: Light and nutrient limitation1. Limnology and Oceanography 21:862-872. Roy, P. S., R. J. Williams, A. R. Jones, I. Yassini, P. J. Gibbs, B. Coates, R. J. West, P. R. Scanes, J. P. Hudson, and S. Nichol. 2001. Structure and function of south-east Australian estuaries. Estuarine, Coastal and Shelf Science 53:351-384. Savage, C., S. F. Thrush, A. M. Lohrer, and J. E. Hewitt. 2012. Ecosystem Services Transcend Boundaries: Estuaries Provide Resource Subsidies and Influence Functional Diversity in Coastal Benthic Communities. PLoS ONE 7:e42708. Seidel, H., C. Löser, A. Zehnsdorf, P. Hoffmann, and R. Schmerold. 2004. Bioremediation Process for Sediments Contaminated by Heavy Metals:  Feasibility Study on a Pilot Scale. Environmental Science & Technology 38:1582-1588. Seitzinger, S. P. 1988. Denitrification in freshwater and coastal marine ecosystems: ecological and geochemical significance. Limnology and Oceanography 33:702-724. Sekovski, I., A. Newton, and W. C. Dennison. 2012. Megacities in the coastal zone: Using a driverpressure-state-impact-response framework to address complex environmental problems. Estuarine, Coastal and Shelf Science 96:48-59. Sharley, D. J., S. M. Sharp, S. Bourgues, and V. J. Pettigrove. 2016. Detecting long-term temporal trends in sediment-bound trace metals from urbanised catchments. Environmental Pollution.

AC C

649 650 651 652 653 654 655 656 657 658 659 660 661 662 663 664 665 666 667 668 669 670 671 672 673 674 675 676 677 678 679 680 681 682 683 684 685 686 687 688 689 690 691 692 693 694 695 696 697 698 699 700 701

15

ACCEPTED MANUSCRIPT

EP

TE D

M AN U

SC

RI PT

Simpson, S., G. Batley, and A. Chariton. 2013. Revision of the ANZECC/ARMCANZ sediment quality guidelines. CSIRO Land and Water Report 8:128. Simpson, S. L., E. J. Maher, and D. F. Jolley. 2004. Processes controlling metal transport and retention as metal-contaminated groundwaters efflux through estuarine sediments. Chemosphere 56:821-831. Spruill, T. B. and J. F. Bratton. 2008. Estimation of groundwater and nutrient fluxes to the Neuse River Estuary, North Carolina. Estuaries and Coasts 31:501-520. Srivastava, D. S. and M. Vellend. 2005. Biodiversity-ecosystem function research: is it relevant to conservation? Annual Review of Ecology, Evolution, and Systematics:267-294. Sundbäck, K. and B. Jönsson. 1988. Microphytobenthic productivity and biomass in sublittoral sediments of a stratified bay, southeastern Kattegat. Journal of Experimental Marine Biology and Ecology 122:63-81. Taylor, S. E., G. F. Birch, and F. Links. 2004. Historical catchment changes and temporal impact on sediment of the receiving basin, Port Jackson, New South Wales. Australian Journal of Earth Sciences 51:233-246. USEPA. 2007. Method 3051A Microwave assisted acid digestion of sediments, sludges and oils. US Environmental Protection Agency, Washington. Vanderborght, J. P., R. Wollas, and G. Bitten. 1977. Kinetic models of diagenesis in disturbed sediments. Part 1. Mass transfer properties and silica diagenesis. Limnology and Oceanography 22:787-793. Viana, P. Z., K. Yin, and K. J. Rockne. 2008. Modeling Active Capping Efficacy. 1. Metal and Organometal Contaminated Sediment Remediation. Environmental Science & Technology 42:8922-8929. von Glasow, R., T. D. Jickells, A. Baklanov, G. R. Carmichael, T. M. Church, L. Gallardo, C. Hughes, M. Kanakidou, P. S. Liss, L. Mee, R. Raine, P. Ramachandran, R. Ramesh, K. Sundseth, U. Tsunogai, M. Uematsu, and T. Zhu. 2013. Megacities and Large Urban Agglomerations in the Coastal Zone: Interactions Between Atmosphere, Land, and Marine Ecosystems. Ambio 42:13-28. Worm, B., E. B. Barbier, N. Beaumont, J. E. Duffy, C. Folke, B. S. Halpern, J. B. Jackson, H. K. Lotze, F. Micheli, and S. R. Palumbi. 2006. Impacts of biodiversity loss on ocean ecosystem services. Science 314:787-790. Zarull, M. A., J. H. Hartig, and G. Krantzberg. 2002. Ecological Benefits of Contaminated Sediment Remediation. Pages 1-18 in G. W. Ware, editor. Reviews of Environmental Contamination and Toxicology: Continuation of Residue Reviews. Springer New York, New York, NY.

AC C

702 703 704 705 706 707 708 709 710 711 712 713 714 715 716 717 718 719 720 721 722 723 724 725 726 727 728 729 730 731 732 733 734 735 736 737

16

ACCEPTED MANUSCRIPT

a) CR

b) GPP

c) NPP

d) NEM

df

MS

P

MS

P

MS

P

Time

3

2.77

0.46

4.24

0.16

4.94

0.16

1 2 2 3 5 2 6 4 5 10 132

16.5 7.30 0.68 0.18 0.56 3.78 2.85 1.16 0.12 1.01 0.66

0.00 0.01 0.79 0.97 0.72 0.03 0.00 0.40 0.98 0.13

37.8 3.04 0.69 3.18 1.48 5.31 1.79 4.68 0.18 1.33 0.30

0.01 0.67 0.68 0.25 0.41 0.31 0.00 0.04 0.97 0.00

24.3 1.22 0.39 3.24 1.49 3.95 2.06 6.64 0.14 1.41 0.35

0.02 0.94 0.83 0.29 0.43 0.60 0.00 0.01 0.99 0.00

P

5.12

0.20

M AN U

10.4 0.86 0.18 2.65 1.30 2.68 2.48 7.26 0.10 1.40

0.06 0.97 0.92 0.43 0.50 0.78 0.00 0.01 0.99 0.00

0.47

EP

TE D

MS

SC

Source

Retention Distance Location (Re) Ti × Re Ti × Di Re × Di Ti × Lo(Re) Lo(Re) × Di Ti x Re × Di Ti × Lo(Re) × Di Residual

RI PT

Table 1. ANOVA of a) community respiration (CR; dark benthic O2 fluxes), b) gross primary productivity (GPP; dark-light benthic O2 fluxes), and c) net primary productivity (NPP; light benthic O2 fluxes), and d) net ecosystem metabolism (NEM; average daily O2 fluxes) between sampling months (Time), water retention characteristic (Retention), location (nested in retention type) (Location (Re)), and distance from stormwater source (Distance). Analyses were run using normalised data and Euclidean distance to construct resemblance matrices. bold = significant at p< 0.05.

AC C

738 739 740 741 742

1

ACCEPTED MANUSCRIPT

NH3 Dark

749 750 751 752 753

NOx Light

PO4 Dark

PO4Light

MS

P

MS

P

MS

P

MS

P

MS

P

MS

P

Time

3

3.32

0.04

0.97

0.84

1.41

0.43

5.22

0.36

3.71

0.04

4.42

0.49

Retention

1

7.48

0.39

0.27

0.65

10.4

0.26

9.18

0.11

0

0.74

2.32

0.37

Distance

2

4.24

0.27

1.89

0.15

1.7

0.08

0.17

0.71

0.5

0.05

2.14

0.01

Location (Re)

2

3.81

0.03

2.75

0.53

4.56

0.09

0.38

0.91

0.86

0.36

3.78

0.54

Ti × Re

3

2.81

0.03

3.44

0.50

2.11

0.29

3.75

0.49

0.62

0.41

1.8

0.78

Ti × Di

5

2.43

0.04

0.78

0.53

0.37

0.83

0.96

0.29

0.21

0.95

0.41

0.89

Re × Di

2

2.57

0.13

1.78

0.15

4.61

0.03

1.14

0.25

0.35

0.49

0.8

0.14

Ti × Lo(Re)

6

0.6

0.54

3.76

0.00

1.33

0.05

4.14

0.00

0.73

0.65

4.86

0.04

Lo(Re) × Di

4

0.9

0.34

0.58

0.64

0.62

0.61

0.38

0.69

0.3

0.89

0.2

0.95

Ti × Re × Di

5

0.61

0.53

0.73

0.57

0.74

0.56

0.76

0.41

1.21

0.43

0.79

0.69

Ti × Lo(Re) × Di

10

0.71

0.44

0.9

0.30

0.91

0.18

0.68

0.17

1.13

0.37

1.32

0.09

Residual

88

0.71

0.75

0.66

M AN U

SC

df

TE D

748

NOx Dark

Source

0.48

1.03

0.49

EP

747

NH3 Light

RI PT

Table 2. ANOVA of dark and light ammonia fluxes (NH3), dark and light NOx fluxes (NOx), and light and dark phosphate fluxes (PO4) between sampling months (Time), water retention characteristic (Retention), location (nested in retention type) (Location (Re)), and distance from stormwater source (Distance). Analyses were run using normalised data and Euclidean distance to construct resemblance matrices. Bold = significant at p< 0.05.

AC C

743 744 745 746

754 755

1

ACCEPTED MANUSCRIPT

SC

Cumul. 0.120 0.145 0.170 0.187 0.200 0.212 0.219 0.224

M AN U

Prop. 0.120 0.026 0.025 0.017 0.013 0.012 0.008 0.005

TE D

P 0.001 0.023 0.022 0.099 0.196 0.219 0.431 0.751

EP

Variable Pseudo-F Cu 12.8 % <63um 2.80 Cr 2.76 Co 1.85 Al 1.44 % TN 1.32 % TOC 0.84 Pb 0.56

RI PT

Table 3. Proportion of variance in benthic O2, NH3, NOx and PO4 fluxes explained by sediment predictor (metals, organic matter enrichment and sediment grain size) variables in forward DistLM sequential tests. (Prop. is the proportion of variance explained by each predictor, Cumul. is the cumulative proportion of variance explained by sequential predictors.) Significant predictors indicated in bold (Figure 4).

AC C

756 757 758 759

2

ACCEPTED MANUSCRIPT 760 761

765 766 767 768 769 770

EP

764

Figure 1. a) Australia with the location of Sydney Harbour indicated by a box, b) Port Jackson estuary and sub-catchments, and c) sampling locations were low retention: Upper Parramatta River (110 km2) and Lane Cove River (80 km2) or high retention: Iron Cove (14 km2); Hen and Chicken Bay (<10 km2). Sampling sites within each location (n=3) are indicated by approximate distance from stormdrain (0, 200, 1000 m respectively).

AC C

763

TE D

M AN U

SC

RI PT

762

1

M AN U

SC

RI PT

ACCEPTED MANUSCRIPT

771

779 780 781 782 783

TE D

light benthic O2 fluxes), c) net primary productivity (light benthic O2 fluxes), and d) net ecosystem metabolism (average daily O2 fluxes) measured in sediment cores collected at increasing distances (0, 200, 1000 m) from a stormdrain in two embayment and two channel locations. Individual bars represent locations in order from left to right; Hen and Chicken Bay, Iron Cove, Upper Parramatta River and Lane Cove River. Measures were averaged over time.

EP

778

Figure 2. a) Community respiration (dark benthic O2 fluxes), b) gross primary productivity (dark-

AC C

772 773 774 775 776 777

784 785 786 2

788 789 790 791 792

AC C

787

EP

TE D

M AN U

SC

RI PT

ACCEPTED MANUSCRIPT

Figure 3. a) Dark and b) light ammonia fluxes (NH3), c) dark and d) light NOx fluxes (NOx) and e)

light and f) dark phosphate fluxes (PO4) measured in sediment cores collected at increasing distances (0, 200, 1000 m) from a stormdrain in two embayment and two channel locations. Individual bars represent locations in order from left to right; Hen and Chicken Bay, Iron Cove, Upper Parramatta River and Lane Cove River. Measures were averaged over time.

793 794 795 3

ACCEPTED MANUSCRIPT

Co

2

%TOC

Cu

SC

Pb Al %TN

M AN U

1

0

Cr

% <63um

797 798 799 800 801 802 803 804

-1 0 1 2 3 dbRDA1 (58.5% of fitted, 13.1% of total variation)

4

AC C

-2

TE D

-1

-2 796

RI PT

3

EP

dbRDA2 (20.4% of fitted, 4.6% of total variation)

4

Figure 4. Constrained dbRDA plot of benthic O2, NH3, NOx and PO4 fluxes fitted to sediment predictor (metals, organic matter enrichment and sediment grain size) variables in forward DistLM sequential tests with AIC selection criteria (Table 3). Lengths of vector overlays indicate the relative influences of fitted predictor variables. Open symbols are embayments and filled symbols are channels. 0 m = diamonds, 200 m = triangles and 1000 m = squares.

4

ACCEPTED MANUSCRIPT

EP

TE D

M AN U

SC

RI PT

Coastal development is a major threat to the estuarine health Urban contaminants enter estuaries and can modify sediment functions We measured metabolism and nutrient fluxes in response to contaminant exposures In situ measures were compared at discharge points between high/low retention Increased O2 consumption was evident closest to poorly flushed discharge points

AC C

• • • • •

1