MEASURING SPATIAL INFLUENCE OF RECYCLED WASTEWATER IN COCKLE CREEK USING STABLE ISOTOPES OF NITROGEN FROM MANGROVES AS PASSIVE AND ACTIVE BIO-INDICATORS

A thesis submitted to the Discipline of Environmental Science and Management, University of Newcastle, in partial fulfillment of the requirements of the Honours Degree of the Bachelor of Science in Marine Science.

David Workman Student ID: 9108473 [email protected]

Submitted February 2018

ABSTRACT

Nutrients form the basic building blocks for living tissues, however excess nutrients of anthropogenic origin are often linked to eutrophication and degraded estuarine ecosystems. Sewage effluent, a key anthropogenic nutrient source, typically has a distinct and elevated δ15N stable isotope signature. As plants and algae grow, the ambient δ15N signature of the nutrient source is generally retained in new tissue, until recycled by the organism or lost in leaf fall, and this signal can be used to identify anthropogenic sources such as effluent discharge or recycled water in an estuary. Despite significant asset investment by Hunter Water in the 1990s diverting all sewage effluent from Lake Macquarie to deep ocean outfall, elevated δ15N signals have been detected at the mouth of Cockle Creek. This study examined the δ15N stable isotope ratios of mangrove leaves and pneumatophore epiphyte communities to distinguish the sources of this elevated signature from several industrial operations within the catchment, and specifically, define the zone of biological influence for Edgeworth WWTP and the recycled water products it distributes. Mangrove leaves and pneumatophores were collected as passive bio-indicators from sites distributed along the banks of Cockle Creek, and pneumatophores were twice deployed as active bio-indicators. Heat maps of δ15N for each indicator, identifying the areas of the creek with relatively high δ15N levels, were used to infer the spatial influence of major nutrient sources. Results from all three methods were comparable, showing a general localised effect of increased δ15N around the WWTP, diminishing to average or lower levels within 300m in all directions. Other elevated sources exist downstream, and the maximum δ15N signal was recorded in the main channel, immediately downstream of the Pasminco rehabilitation site near Boolaroo. Some sites were depleted in δ15N relative to other sites, which was associated with mangrove marsh chemistry at some locations. For more urban sites, other input sources with a lower δ15N signature may be diluting observed δ15N. Mangrove leaves and epiphytes could distinguish differences in the δ15N signature between sites, confirming their suitability as passive indicators of nutrient sources. Active epiphyte deployments distinguished elevated δ15N sources under spatial analysis through variations in enrichment, but did not fully acclimatise with ambient surroundings after 5 days. A 14-day deployment is recommended. With suitable time to acclimate, the method shows strong potential for identification of sewage-derived nutrient sources in an estuary.

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Academic Supervisors Dr Troy Gaston Senior Lecturer in Environmental Science School of Environmental and Life Sciences Faculty of Science and IT E: [email protected]

Dr Vincent Raoult Postdoctoral fellow School of Environmental and Life Sciences Faculty of Science and IT E: [email protected]

ACKNOWLEDGEMENTS

This project has been funded through a Water Research Australia scholarship grant (Project # 4110 -17), sponsored by Hunter Water. Thanks to Carolyn Bellamy and the WaterRA team for your wonderful support.

The University of Newcastle provided boat and lab facilities, boat drivers and field assistants, and the support and encouragement of the whole faculty has been vital. Special thanks are due to Troy and Vincent, for their patience and understanding, as they offered suggestions, insights, critiques and feedback along the way. Thanks also to Tim Smith, Dan Creevey and again, Vincent for assisting with boat driving and fieldwork. Tom Ryan’s guidance and advice in the lab were very helpful, and his eclectic musical tastes made the lab a fun and interesting place to be.

Thanks also to Emma Graham and Symon Walpole of Lake Macquarie City Council , who provided access to Council’s research library, and data from the Lake Macquarie monitoring program.

To my sons, Sam and Ben, and indeed all my family, thank you. Your interest, encouragement and enthusiasm along the way have been vital. To the friends who’ve forgotten what I look like, thanks for understanding when I couldn’t make it to whatever events I missed.

And finally, to Kristy, without whom none of this is possible, words could never express my gratitude or love, but thank you for everything.

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Table of Contents Abstract ...... 1

Acknowledgements ...... 2

Introduction ...... 5

Background ...... 5

Sewage in estuarine systems ...... 6

Bio-indicators – Active and Passive indicators ...... 7

Tracking sewage using stable isotopes - fractionation, depletion and enrichment ...... 7

Using mangroves and epiphytes as bio-indicators ...... 9

Impacts of varying nutrient supply & water quality ...... 11

Conceptual model of relevant nutrient sources and processes ...... 12

Lake Macquarie & Cockle Creek ...... 14

Past issues and nutrient history...... 16

The problem ...... 18

Materials & Methods ...... 21

Site selection ...... 21

Rainfall ...... 22

Passive bio-indicator sampling ...... 23

Mangrove leaves ...... 23

Epiphytes from pneumatophores ...... 23

Active bio-indicator deployment ...... 23

Laboratory analyses ...... 24

Data Analysis ...... 25

Spatial Analysis ...... 26

Results ...... 27

Passive Bioindicators ...... 27

Isotope data in mangrove leaves ...... 27

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Isotope data in passive epiphytes ...... 30

Spatial results in passive bioindicators ...... 31

Mangrove leaves ...... 31

Epiphytes (Passive) ...... 32

Active bio-indicators ...... 33

Reference data for active deployment of epiphytes ...... 33

Isotope data for active bioindicators ...... 33

Spatial results for active bioindicators ...... 34

C:N ratios ...... 36

Mangrove leaves ...... 36

Epiphytes (passive) ...... 36

Active bioindicators ...... 36

Water quality ...... 37

Discussion ...... 41

Passive Bio-indicators ...... 42

Active bio-indicators ...... 44

Potential inputs and sources ...... 45

Water quality parameters ...... 49

Conclusion ...... 51

References ...... 52

Appendices ...... 57

Appendix A – Comparison of pilot study & August results ...... 57

Appendix B – Miscellaneous Data ...... 60

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INTRODUCTION

BACKGROUND

Estuaries are permanently or periodically open water bodies connected to the sea that exhibit variations in salinity, turbidity, pH and dissolved oxygen related to oceanic mixing (Day 1980). Australia has high seasonal and spatial variability in rainfall and run-off, which leads to a broad range of estuary types and conditions, both temporally and spatially (Roy et al. 2001). Estuaries are typically highly productive systems (Deeley & Paling 1999), supporting extensive biotic communities and commercially important fishery species (Roy et al. 2001, Sheaves et al. 2012, Creighton et al. 2015), but they are also subject to increased nutrient, sediment and pollutant loads as a result of urban development being highly concentrated on Australia’s coastal fringe (Deeley & Paling 1999).

Biologically available nitrogen supplied in the form of ammonia, nitrite or nitrate, forms one of the key nutrients for plants and other autotrophs to create new tissue. Its availability usually limits primary production in estuarine systems in most temperate regions (Carpenter et al. 1998, Roy et al. 2001, Scanes et al. 2007). When present, it is either rapidly taken up in photosynthetic activity and assimilated into aquatic plants and algae, or settles quickly into bottom sediments (Leonard & Steven 2001, Merriam et al. 2002, Scanes et al. 2007, Tan et al. 2013). Organic carbon in the form of dead plant and animal tissue or detritus is another key requirement for growth in plants (Boon & Bunn 1994, Vizzini & Mazzola 2003, Gaston et al. 2004, Zhen & Zhu 2016, Signa et al. 2017). Although most plants and algae draw carbon dioxide from the air in photosynthesis, not all plant requirements can be met this way, and organic carbon - in soils, detritus or dissolved form - is required.

In excess, however, nutrients can be detrimental. The most commonly associated effect is an algal bloom, where phytoplankton biomass increases due to elevated nutrient availability. The ensuing increased biological demand often leads to a state of eutrophication with reduced oxygen levels, fauna mortality, and toxic by-products (Carpenter et al. 1998). Storm events and associated run-off that carry leaf litter and detritus from the catchment are the most common natural processes allowing nutrients to enter a system (Scanes et al. 1995, Roy et al. 2001, Merriam et al. 2002). Run-off from urban sources – potentially carrying sewage,

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fertilizer, garden clippings, and agricultural and industrial waste - is also significant (Table 1, Carpenter et al. 1998, Nixon 2009).

Table 1: Anthropogenic nutrient sources common in developed catchments (Carpenter et al., 1998) Point Sources Non-point sources Wastewater effluent Runoff – agricultural (incl return flows from irrigated agriculture) Runoff/leachate - waste disposal sites Runoff – pasture and range Runoff - animal feedlots Urban runoff – unsewered areas and sewered where pop’n < 100,000 Runoff - mines, oilfields Septic tank leachate, runoff – failed septic systems Sewer storm outfalls (pop’n > 100,000) Runoff – construction sites < 2ha Overflows of combined sanitary/storm sewers Runoff – abandoned mines Runoff – construction sites > 2ha Atmospheric deposition over water surface Runoff – unsewered industrial sites Activities on land that generate contaminants – eg logging, wetland conversion, construction and development of land or waterways

SEWAGE IN ESTUARINE SYSTEMS

Although agriculture is frequently the main anthropogenic nutrient source (Carpenter et al. 1998, Quiggin 2001, Azad & Ancev 2010), human effluent is becoming increasingly recognized as a problem (Scanes et al. 1995, Divers et al. 2014). An estimated 30-50% of nutrients within the near shore coastal region surrounding Sydney are sewage derived (Gaston et al. 2004). The impacts of sewage in estuarine ecosystems are well understood, and mostly related to algal blooms and eutrophication due to excess nutrients (Costanzo et al. 2001, Keuskamp 2003, Nixon 2009, Tan et al. 2013, Lequerica & McInnes 2016). Other impacts of sewage include: degraded water quality with human and livestock health effects, loss of biodiversity (including important commercial and recreational species), shellfish poisoning and industry impacts, loss of aquatic plants and reefs, increased weed and nuisance species, changes to marine food webs, and increased odours (Scanes et al. 1995, Carpenter et al. 1998, Quiggin 2001, Nixon 2009).

Due to leaking pipes and infrastructure, sewage often enters a system accidentally, and with emerging planning and design trends that encourage the use of recycled water for industrial and agricultural uses (Lim et al. 2010), the risk of sewage-derived nutrients entering a system increases. Current monitoring programs based on water quality parameters, nutrient loads and species assemblages can only present a snapshot of conditions at the time of measurement, and their spatial and temporal variability make them unreliable indicators. Furthermore, water

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quality parameters are of little benefit in determining nutrient sources, or the biological fate of these nutrients (Scanes et al. 2007).

BIO-INDICATORS – ACTIVE AND PASSIVE INDICATORS

One approach to overcome this uncertainty is the use of bio-indicators, a diverse range of biological organisms that accumulate contaminants in their tissues (Moore 2003), or provide information through changes to physical structure or chemical content, variations in abundance, or their presence or absence. Useful bio-indicators require particular characteristics, namely the capacity to accumulate contaminants without being lethal, they must be sedentary organisms to avoid losses, and be representative of the area. They must also be abundant in the area of interest, easily identified and sampled, hardy and provide enough tissue for analysis (Phillips & Rainbow 1994).

Bio-indicators can be passive or active. Passive bio-indicators are found naturally at the site of interest, and can give a picture of cumulative or average conditions over the organism’s life. The temporal scale represented is related to the tissue turnover rate – how long it takes to recycle nutrients from dead or dying tissues (Cifuentes et al. 1996, Fry et al. 2000, Costanzo et al. 2001, Alongi et al. 2005a, Fry 2006). In contrast, active bio-indicators are transplanted from an area of low pollution to the study area, to assess any changes during a specific deployment period, and infer ambient conditions as well incorporation rates during the study period (Moore 2003). In addition, from a practical perspective, they can be deployed in locations where the monitor species does not naturally occur.

TRACKING SEWAGE USING STABLE ISOTOPES - FRACTIONATION, DEPLETION AND

ENRICHMENT

Stable isotopes in bio-indicator species can determine sewage-derived nutrients when traditional ecological methods may not. For example, Scanes et al. (1995) found no impact on species assemblages from deep-water ocean outfalls, whereas Gaston et al. (2004) were able to demonstrate using stable isotopes that 30-50% of local nutrients were sewage-derived. Morrissey et al. (2013) found a link in effluent-treated waters between stable isotope enrichment and macroinvertebrate indices, common indicators of freshwater ecosystem health (Azad & Ancev 2010).

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Most elements occur as more than one different stable or non-radioactive isotope, where the number of neutrons in the nucleus may be 1 or 2 more or less than ‘usual’. Nitrogen occurs naturally with an extra neutron (15N), which is stable, and while it makes up a minor component of overall nitrogen, the isotopic ratio of 15N to 14N in naturally sourced nitrogen is generally universal (Fry 2006). Marginally lighter but otherwise indistinguishable, 14N is preferred for some reactions, especially those that are bacteria- and enzyme-moderated (Fry 2006, Rush 2003), altering the ratio in both the products (depleted in 15N) and the remaining pool (left enriched in 15N). Through analysis of this 15N :14N ratio, and comparing it to known source ratios, the source of nitrogen in an ecosystem can be determined.

Untreated sewage is raw human waste, where 14N is preferentially excreted, leading to a depleted 15N signal (Fry 2006). Soon after excretion urea is hydrolysed, causing some fractionation, as does volatilization of ammonia subsequently formed (Costanzo et al. 2001). Consequently, the remaining pool of ammonium is enriched in 15N. Conversion to nitrates (nitrification), and subsequent de-nitrification to nitrogen gas is also accompanied by fractionation (Costanzo et al. 2001, Fry & Cormier 2011). Furthermore, in sewage treatment processes bacteria preferentially take up 14N, leading to treated effluent output being highly enriched in 15N (Fry 2006). When released into an oligotrophic waterway macrophytes and primary producers usually take up all nutrients as they become available, including 15N- nitrogen, and the isotopic signature is incorporated into new tissue (Fry et al. 2000, Costanzo et al. 2001, Costanzo et al. 2005, Fry 2006). The subsequently depleted or enriched isotope signature can be detected until the tissue is recycled, or lost as leaf litter, fallen branches or detritus (Fry 2006).

Various isotope approaches that mainly focused on passive bio-indicators have been used to detect anthropogenic inputs, with some practical considerations that limit their widespread use. For example, using water samples for nutrients in the water column (Divers et al. 2014, Zhen & Zhu 2016) requires extensive analytical chemistry techniques, while sampling logistics are difficult, and results can be highly variable. Another approach is to use animal tissues as passive bio-indicators. Morrissey et al. (2013) measured enrichment in macroinvertebrates, and Gaston et al. (2004) demonstrated sewage-derived nutrient uptake into the marine food web. The differing isotopic signatures for each species, and for each tissue type, give an indication of time-scale of exposure based on the nutrient dynamics particular to that tissue type (Gaston et al. 2004, Mohan et al. 2016). Uncertainty regarding

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individual tissue responses usually requires laboratory diet tests to determine or confirm local responses (Bouillon et al. 2008), while the requirement for ethics approval, and logistics of controlling for animal migration further complicate the use of animals as bio-indicators.

Plants have been used for stable isotope analysis to bypass some of the limitations of animal tissues. In Hawaii, the Red Mangrove (Rhizophora mangle Linnaeus) was used as a passive bio-indicator to track groundwater sources in coastal areas, and demonstrate increased anthropogenic input from the watersheds of developed areas (Fry & Cormier 2011). Macroalgae have also been used as active bio-indicators (Costanzo et al. 2001, Costanzo et al. 2005), where cultures of Catenella nipae, an epiphytic algal species, were deployed in Moreton Bay, Brisbane. By analysing the tissue grown in situ over the five-day sampling period, a map was developed showing the contribution of sewage-derived nutrients in the bay. The main benefit of this active deployment was that it allowed measurements where naturally occurring plants and algae had disappeared (Costanzo et al. 2001). These methods are adapted in the present study.

Carbon can be analysed in a similar manner to nitrogen, but while δ13C (13C:12C) can sometimes give an indication of sewage loading into a waterway, the results can be heavily confounded by carbon sources of varying δ13C, unrelated to sewage, within the catchment (Cifuentes et al. 1996, Gaston et al. 2004, Fry 2006, Zhen & Zhu 2016). Ambiguities also arise due to uncertainties about carbon cycling in mangrove systems, dissolved inorganic carbon (DIC) dynamics, and the relative importance of imported carbon (Bouillon et al. 2008). In estuaries, and mangroves in particular, δ13C can be increased by high salinity or growth stress (Fry & Cormier 2011), and as a consequence, a clear decline in 13C enrichment can usually be seen when moving from a fully marine environment upstream towards a more terrestrial environment (Bouillon et al. 2008). These factors can limit the potential for δ13C analysis in pollution studies.

USING MANGROVES AND EPIPHYTES AS BIO-INDICATORS

In Cockle Creek as with many temperate and tropical estuaries, mangroves are ubiquitous along most of the banks and foreshore, extending as far as the tidal influence (Alongi et al. 2005a). Their abundance makes them an ideal bio-indicator, particularly where species used

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by previous authors, such as oysters (Moore 2003), seagrass and seagrass epiphytes (Clarke 2016), or particular algal species (Costanzo et al. 2001, Rush 2003), are not present estuary- wide. Mangroves have been widely used in stable isotope analyses (Daigle et al. 1990, Cifuentes et al. 1996, Fry et al. 2000, Alongi et al. 2005b, Bouillon et al. 2008, Kousbrouek et al. 2017). They are capable of highly productive and efficient nutrient assimilation (Alongi et al. 2005a), making them an ideal and readily available passive bio-indicator of predominant nutrient conditions over the life of the tissues sampled.

Estimates of the leaf lifespan vary between 13.7 months (Ellison 2002) to less than 2 years (Alongi et al. 2005a) and up to 3 years (Kumar et al. 2011). This life span is influenced by temperature, rainfall, and by the import or export of nutrients or organic material through tidal exchange, riverine flooding and wave action (Signa et al. 2017). Sampling leaves is simple and cost-effective, and Kousbrouek et al. (2017) demonstrated low variability in 15N between leaf replicates from the same site. No marked 15N differences have been found between green and senescent leaves, but there are some differences between tissue type, where pneumatophores can be enriched relative to the leaves (Bouillon et al. 2008).

The roots and pneumatophores (emergent roots standing out of the water column) of mangroves form a natural habitat for algae and other epiphytes (Runcie et al. 2003). These epiphytes form a significant part of estuarine food web dynamics and nutrient cycles, and are dominated by algal mats and cyanobacteria (Rodriguez & Stoner 1990). Variation in epiphyte communities occurs on spatial and temporal scales, and can reflect nutrient source variation (Boon & Bunn 1994, Melville & Connolly 2003). The subsequent effect on δ15N of this variation in species composition is unknown. Cyanobacteria can fix atmospheric nitrogen, while red algae have high rates of photosynthesis (Cifuentes et al. 1996). These processes could lead to dilution of any elevated signature with atmospheric nitrogen, or increased fractionation through photosynthetic enzyme selection on fixation. Responses from epiphytes in suspended particulate matter are influenced more by these processes than from variation in terrestrial inputs (Cifuentes et al. 1996), but phytoplankton generally assume the isotopic signature of the ambient dissolved ammonium content (Mariotti et al. 1984). Clarke (2017) used seagrass epiphytes to successfully distinguish different nutrient sources in Lake Macquarie.

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The temporal variation of isotope analysis in mangrove epiphytes is dictated by the tissue turnover rates of the epiphyte community. Rather than averaged over several months or years, the epiphyte tissues are recycled within days or weeks. As passive bio-indicators collected in situ, epiphytes can provide a resolution of nutrient sources over this more immediate timeframe. The main limitation of this method is it only provides information on nutrients reaching the banks of the creek where mangroves occur.

The shorter turnover periods for epiphytes make them ideal for deployment as active bio- indicators (Costanzo et al. 2001). Deployment of active bio-indicators at chosen locations allows observation of sewage-derived nutrients in the water column, rather than at passive locations on the banks. Such deployment requires sufficient time for tissue turnover, so that all old epiphyte tissue will be replaced, and new tissue can assume the isotopic signature of the ambient nutrients. Costanzo et al. (2001) found a turnover rate of 5 days for C. nipae, a more stable indicator than Sea Lettuce (Ulva lactuca), which quickly acclimates with ambient conditions (Runcie et al. 2003). Working with Microdictyon umbilicatum, another common species in NSW, Rush (2003) found tissue turnover within 7-10 days, both in laboratory conditions, and in situ in the Hawkesbury River, NSW.

Physical parameters are also important considerations for stable isotope analysis of epiphyte communities. Temporal shifts on a seasonal scale can influence isotope composition of macroalgae and phytoplankton (Daigle et al. 1990), which can be mirrored by variations in some higher trophic levels (Vizzini & Mazzola 2003). Spatial trends in enrichment based on hydrodynamics and nutrient availability have also been observed (Guest et al. 2004), though not significant compared to variation among sites. Due to the lower water velocities in boundary layers around epiphytes, they are generally enriched compared to mangroves and phytoplankton, (Daigle et al. 1990). Phytoplankton in the water column, by contrast, is opportunistic and may be relatively depleted in line with ambient water conditions (Daigle et al. 1990, Bouillon et al. 2008). These factors must all be considered as potential sources of variation in isotope results from epiphytes.

IMPACTS OF VARYING NUTRIENT SUPPLY & WATER QUALITY

Although a very useful method to determine nutrient sources, stable isotope analysis relies on ammonium or nitrate uptake as the rate-limiting step, and all nutrients being effectively

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consumed (Mariotti et al. 1984, Fry 2006). Although in oligotrophic estuaries, excess nutrients usually only occur after rainfall driven catchment drainage (Day 1980, Roy et al. 2001), during these periods, other factors may become important.

Fractionation changes may be hard to predict quantitatively based on water quality parameters, but broad scale relative changes can be predicted (Jennings & Warr 2003), and thus monitoring of physicochemical parameters can eliminate some potential confounding factors. Nutrient pathways are strongly regulated by pH for example, thus any change in pH can alter fractionation outcomes (Rush 2003). Other water quality parameters can also regulate these reaction pathways, however light reduced to half of natural levels, or temperature and salinity within naturally encountered ranges, had no effect in the Hawkesbury River (Rush 2003). Very shallow water can cause changes in fractionation due to increased evaporation, which is highly selective for 14N, leaving the pool enriched (Rush 2003, Fry 2006).

High carbon:nitrogen ratios may further indicate a switch to lipid production (Logan et al. 2008), due to excess or alternate nutrients, with changes to nutrient assimilation pathways and fractionation rates (Macko et al. 1987, Yamaguchi et al. 2017), and especially those pathways mediated by enzymes (Rush 2003). Nutrients as ammonium, nitrate or nitrite ions also have varying fractionation rates (Rush 2003), and nitrification and denitrification due to excess nutrients also increases fractionation (Mariotti et al. 1984, Fry & Cormier 2011).

CONCEPTUAL MODEL OF RELEVANT NUTRIENT SOURCES AND PROCESSES

The interpretation of any stable isotope analysis requires not only an understanding of source signatures, but also consideration of the nutrient pathways and processes detailed above that may have altered these signatures (Rush 2003). Within any estuarine system such as Cockle Creek in the Lake Macquarie system, there are multiple nutrient inputs as well as processes to be considered (Figure 1) including: point (for example, wastewater) and non-point (for example, diffuse run-off), anthropogenic (eg agriculture) and natural sources (eg atmospheric deposition).

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Figure Figure 1: Conceptual Conceptual 1: model of potential nitrogen sources and the factors affecting δ 15 N

enrichment for mangroves and

epiphytic

algaeCockle in Creek

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LAKE MACQUARIE & COCKLE CREEK

Lake Macquarie is the largest coastal lake in Australia (area 110.6 km2), draining a catchment area of 680.8 km2. It is approximately 24km long and 3.2 km wide in a roughly north south orientation (Figure 2). It has a large volume relative to other coastal lakes of NSW, of 9 x 108 m3 (or 900,000ML). Depth in the lake averages 6-8m, with a maximum of around 11-12m. Mean annual rainfall in the catchment is 1230.5mm, and the local population was estimated at 200,000 in 2011 (AWACS 1995, LMCC 2017).

The lake has a restricted channel, low fluvial inputs (2% of tidal exchange), and minimal tidal intrusion, with a small tidal prism of between 3.5 and 5%. Full oceanic exchange occurs approximately 2-3 times per year (AWACS 1995, Watterson et al. 2010). The peak tidal range is generally less than 30cm, but the lake water level is estimated to be rising approximately 0.12% per annum, possibly increasing the tidal prism (AWACS 1995, Watterson et al. 2010). The lake was subject to frequent natural closing of the entrance in the past, but with regular dredging, the entrance has remained permanently open (Roy & Crawford 1984, AWACS 1995, Watterson et al. 2010). Large storms and associated sea level changes can have greater impact than tides on the lake water level, and are also the typical pathway for nutrients to enter the lake (AWACS 1995). There are no ‘short circuits’ from any of the riverine inputs to the outlet channel (ERM Australia 2000).

Cockle Creek (32.96233 S 151.61443 E, Figure 2) is the second largest sub-catchment in Lake Macquarie, up to 17% of the total lake catchment land area. Within the sub-catchment, land area is 46 times the water area, compared to 6:1 for the whole of Lake Macquarie, and subsequently Cockle Creek supplies about 44% of the sediment yield and a disproportionate amount of the anthropogenic inputs for the lake (AWACS 1995, MHL 2002). The two upstream weirs - at Weir Rd, Cockle Creek, and at Bower Oval on Winding Creek (Figure 3) - act to prevent tidal intrusion into catchment during low flow, but mainly slow run-off into the creeks during flood events (MHL 2002). Further, a sill of marine derived material has formed at the mouth of Cockle Creek, which prevents full flushing of the creek bottom waters (AWACS 1995, MHL 2002). The upper catchment is relatively undeveloped, with some reducing coalmine activity and up to 80% natural catchment vegetation (Hodgson 2003). However, the lower catchment includes heavily developed areas like Cardiff, Glendale, Charlestown, Edgeworth and Cameron Park.

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Figure 2: Cockle Creek catchment area outlined in red, with the Cockle Creek estuary shown in blue (Google Earth Pro, 2017). The catchment area is large compared to the estuary size, with a wide variety of land uses within it. Inset shows the location of Cockle Creek at the northern end of Lake Macquarie.

Tidal Weir

WWTP

Tidal Weir Golf Course Overflow dam

Coal washery Pasminco (1897 - 2003) Rehabilitation site

Washery & dam – both overflow into creek

Cockle Bay

LAKE MACQUARIE

Figure 3: Main features of the Cockle Creek lower catchment, adapted from Google Maps, 2017.

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PAST ISSUES AND NUTRIENT HISTORY.

Lake Macquarie and Cockle Creek in particular has had high levels of nutrients reported for decades, most notably sewage-derived nutrients (AWACS 1995, ERM Australia 2000, Hodgson 2003). Although a popular swimming spot in the 1920s with clear, sandy-bottomed water over two metres deep, by the 1940’s silt from mine workings in the upper catchment, power station ash dumping, and effluent from concentrated greyhound breeding operations had led to environmental pollution in Cockle Creek (LMC Library 2017b).

The Sulfide Corporation lead smelter had opened on the eastern bank of Cockle Creek in 1897 (Figure 3, Pasminco site), and over its life, the plant underwent many uses, including fertilizer production from 1969 until 2009 (Manidis Roberts 2010). Fertilizers are typically very high in nitrogen, but as the usual method of production involves fixation of atmospheric nitrogen, it will generally have a δ15N close to zero (Fry 2006).

In 1920, Waratah Golf Club commenced preparing a 102 acre (41.3 ha) site with an extensive 2.5 km frontage to the eastern side of Cockle, Brush and Winding Creeks (Figure 3). Large tracts of riparian vegetation were cleared to make way for heavily fertilized and cultivated fairways and greens, and fresh manure was applied. When Edgeworth Wastewater Treatment Plant (WWTP) was opened in 1959, followed by West Wallsend in 1963, both discharged effluent outflow into Cockle Creek (LMC Library 2017c, a, d). In 1982, following a century of mining activity, the former Stockton Borehole Colliery developed what became the Oceanic Coal Washery, directly across the creek from the southern end of the golf course (Figure 3). Now closed since mid 2016, several adjacent storage dams remain, and washery- derived waste, in times of high rainfall, overflows into Cockle Creek (Clibborn 2017).

Ecological studies have confirmed the impacts of this industrial history for Cockle Creek. Deoxygenation, lower pH than the body of the lake, and large pH fluctuations attributed to algal blooms associated with effluent discharge are some of the issues known to affect this creek. These findings were supported by consistently high chlorophyll-a readings over the same periods, and they all support conclusions of elevated nutrient history (AWACS 1995). Fauna studies indicate signs of environmental stress at the mouth of Cockle Creek (AWACS 1995, MHL 2002), and annual mean organic nitrogen (ammonia) concentrations were frequently measured above ANZECC guidelines. Nutrient levels were noted to be high near

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the washery overflow in several older studies(AWACS 1995, ERM Australia 2000, Hodgson 2003). Cockle Creek had significantly higher levels of faecal coliform bacteria than the rest of the lake between 1983 and 1994, and regularly failed to meet ANZECC guidelines for primary and secondary contact. It has also, at various times, been impacted by nuisance algal blooms (AWACS 1995).

With sewage seen as a significant source of these problems, Hunter Water spent $310 million on WWTP upgrades in the 1990s, removing all WWTP outfalls from Lake Macquarie (Hunter Water 2011), including West Wallsend and Edgeworth from Cockle Creek. Deep water outfalls were found to be more sustainable, having negligible ecological effects, and none in- shore (Scanes & Philip 1995). Some recent studies indicate Cockle Creek and Cocked Hat Creek in particular still have high nutrient loads and suspended solids, with storm water thought primarily responsible (Umwelt 2015), while recent monitoring data from Lake Macquarie City Council (LMCC) indicates nutrients are generally excessive only after rainfall (Appendix B).

Edgeworth WWTP, at the junction of Cockle and Cocked Hat creeks, provides services to Charlestown, Hillsborough, Cardiff, Boolaroo, Speers Point, Glendale, Edgeworth, Barnsley and Killingworth. It has the capacity to treat sewage from a population of up to 70,000. Incoming sewage receives screening, de-gritting, flow measurement and odour control at the inlet, followed by biological treatment through a Modified Ludzack-Ettinger (MLE) process (Hunter Water 2011). Sewage is then mixed with activated sludge, recycled solids from later in the treatment process. The sludge contains live bacterial communities that provide some disinfection, and remove nitrogen through denitrification, with strong fractionation occurring at this point (Costanzo et al. 2001, Fry 2006, Bouillon et al. 2008). Effluent output is then passed to 2 secondary clarifiers, where the sludge can be recycled as activated sludge to be mixed with the effluent inflow, or allowed to settle in the clarifying ponds. Further fractionation due to evaporation, continued microbial activity and denitrification occurs here also (Costanzo et al. 2001, Fry 2006, Fry & Cormier 2011). Settled sludge is separated and de-watered, before use mainly in coalmine site rehabilitation. Water is extracted from the ponds, then undergoes UV exposure before storage in tanks, ready for commercial supply, or on-site irrigation at the WWTP (Hunter Water 2011). Excess water is sent to the effluent settlement pond, before pumping to Toronto, then on to ocean outfall. A schematic is shown at Appendix B.

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Recycled water provision is a key component of Hunter Water Recycling & Reuse initiatives. Little research has been done to delineate the influence of these recycled products in local aquatic ecosystems, but guidelines for best practice suggest they should be carefully managed (including buffer zones) when used near surface waters (DEC 2004). Newstan Oceanic Coal Washery (prior to its closure) was the main recycled water user (457 ML pa to May 2016), with Waratah Golf Club using 97 KL in the same period (McKenzie 2017). Excess wastewater is pumped via Toronto to Belmont WWTP, then to the ocean outfall (EPA 2017). Licensed for up to 30,000 ML of discharge annually, the plant currently treats around 6000ML annually or 20% of capacity (Hunter Water 2011).

THE PROBLEM

With the significant capital outlays spent on eliminating effluent discharge in Lake Macquarie, sewage should not be expected to be a significant nutrient source in the lake or any of its tributaries. However, recent studies have shown the presence of elevated 15N in seagrass at the mouth of Cockle Creek (Figure 4, Clarke 2016). Since no readings were taken within Cockle Creek itself, the source of these elevated signatures is not clear. Edgeworth WWTP discharges all effluent via ocean outfall, but broken or leaking assets could be allowing effluent to escape. Although the use of recycled water has decreased without coal washery operations, the golf course is still a significant user. Recent rehabilitation at the Pasminco site has mobilised large amounts of sediment, and recycled water is used in the rehabilitation works. On-site sewage systems are also present within the catchment, and could be contributing to nutrient inputs.

The purpose of this study is to determine the source(s) of this elevated signature. Specifically, this study seeks to resolve whether the use of recycled water within the Cockle Creek catchment is contributing to the observed 15N results. Using stable isotope analysis of nitrogen and carbon from mangrove leaves and epiphytes, the study aims to illustrate the recent and medium-term history of nutrient sources in Cockle Creek. A secondary aim was to define the spatial range - the zone of influence - of sewage-derived nutrients from recycled wastewater.

The central hypothesis of this study is that mangrove tissues will exhibit high δ15N in locations where anthropogenic (namely sewage-derived) inputs are present. A further

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hypothesis is that these elevated signals will be located near the WWTP, and possibly at other locations, including the golf course and Pasminco rehabilitation site, both recycled water users. Interpolated maps showing the extent of biological influence of sewage derived nutrients will be presented as per Figure 4, featuring visual modelling of areas with high δ15N (and δ13C) signals, relative to their surrounding environment.

The significance of this study is the potential to develop and refine a novel, simple and effective method to detect sewage-derived nutrient sources. At the same time, it will provide much-needed information on the impacts of recycled water application near an estuary. Both of these outcomes will be of great value for estuary users and managers, WWTP operators and recycled water users.

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Speers Point

Dora Creek reference site

Figure 4: Elevated 15N signal in the northern end of Lake Macquarie (circled in red), and at Speers Point in particular. Dora Creek reference site is marked in red (Clarke, 2016).

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MATERIALS & METHODS

SITE SELECTION

For mangrove leaf and epiphyte passive sampling, 16 pilot sites were selected at locations along Cockle Creek, Winding Creek and Cocked Hat Creek (Figure 5; Sites 1-15 and 4a). These sites were chosen with reference to the location of tributary streams, creek junctions, storm inlets, potential nutrient sources, and the availability of mangroves. A site was located at the Barnsley Weir – the tidal and saline extent of Cockle Creek (Site 1), and at the navigable extent of each of the tributaries (Sites 4 and 5). Deployment of active bioindicators was also based on these site locations. Further passive leaf sampling (August 30 2017) was conducted at 11 of the original sites (Sites 2, 3, 4a, 6-12 and 14), and an additional 14 sites spaced between them (Figure 5, Sites A-N), to provide finer spatial resolution of leaf δ15N.

Figure 5: Cockle Creek study sites in relation to key catchment features.

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RAINFALL

Daily rainfall (Figure 6) from Barnsley monitoring station (Manly Hydraulics Laboratory, MHL) and Edgeworth WWTP (Bureau of Meteorology rainfall station) were compared with water levels from MHL water level gauges at the Barnsley Weir and Cockle Creek Railway Station, (Figure 6) for the period before and throughout the study. Generally, the study period was a dry year, with infrequent rain in occasional large pulses. Rainfall on the 31st March 2017 (47mm), and other falls throughout the month, took the monthly total to 275mm, more than double the all-time monthly average (118mm, Appendix B). June 2017 had 10 days of rain, and four days above 20mm. In contrast, May had just 6 days of rainfall, with 14mm on 22nd May the only day above 3mm. In all other months from November 2016, monthly rainfall was below the all-time average for that month. While rain events raised the water level at the weir, there were rarely any observable differences to levels at Cockle Creek Station, which were dominated by the small (-0.1 – 0.4 m) but regular semi-diurnal tides. Even at the weir, flood effects were quick to dissipate when rainfall ceased.

1st mangrove passive sampling 1/2/17 for Active deployments leaves and 1 (1-6/5/17) & 2 (23- pneumatophores 30/6/17) 2nd mangrove leaf sampling 30/8/17

Figure 6: Daily rainfall from a MHL station at Barnsley, and Edgeworth WWTP (BOM), and water levels at the weir and railway station, over the 13 months from October 2016 to October 2017 (BOM 2017).

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PASSIVE BIO-INDICATOR SAMPLING

Mangrove leaves The first passive bio-indicator sampling event, as a pilot study for mangrove leaves, took place on February 1 2017. A second passive sampling of leaves, with additional sites (A-N), occurred on August 30 2017. Mangrove leaves were taken from individual trees within 20m of the site coordinates, at approximately 1.5 – 1.8 metres above water level where possible. Sampled leaves were selected on the basis of having no obvious signs of decay or damage. Leaves from each site were placed into individually labelled zip lock bags, and placed on ice during transport to the lab, where they were frozen until processing.

Epiphytes from pneumatophores

Also at the first passive bio-indicator sampling event (February 1 2017), five mangrove pneumatophores, approximately 10-15cm long, were harvested from all sites as passive indicators. Taken from distinct locations within 20m of each site, they were placed into ziplock bags on ice for transport to the lab, where they were frozen until processing.

ACTIVE BIO-INDICATOR DEPLOYMENT

Active bio-indicator sampling consisted of two deployments of mangrove pneumatophores in 2017: 5-11 May (dry weather conditions) and 23-30 June (following wet weather - 67.5mm total on 7 of the previous 14 days, and 121mm since 1 June). Clear 500mL jars were perforated with approximately 18 x 1cm holes spaced evenly around their circumference to allow free passage of water and nutrients, but minimise interactions with larger herbivores. For each site, five jars were cable-tied to a float apparatus via their screw-on lids (Figure 7). Pneumatophores were harvested from a reference site at Dora Creek known not to have elevated signature for nitrogen (Clarke 2016). One pneumatophore was placed randomly into each jar, attached to the float and deployed at each site for 6-7 days. At the end of the deployment period, pneumatophores were placed into labelled ziplock bags, and placed on ice for transport to the lab where they were kept frozen until processed.

Samples of final effluent (n = 3; post UV treatment before distribution as recycled water) were collected from Edgeworth Wastewater Treatment plant on August 29 2017. Samples were collected in 1L HDPE bottles and placed on ice for transport to the laboratory.

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Figure 7: Active monitoring apparatus - each algal trap has a pneumatophore inside the jar.

On each sampling occasion, water quality data was recorded at every site using a Hydrolab DS 5X probe, including temperature (oC), specific conductivity (mS/cm), pH (pH units), turbidity (NTU), dissolved oxygen (% and g/L), chlorophyll A (ug/L) and salinity (ppt). Due to calibration errors, pH recordings were exceptionally high for all sites in May and June and not used further. Dissolved oxygen readings were not obtained at occasional sites due to malfunction of the probe.

LABORATORY ANALYSES

Mangrove leaves were thawed, rinsed with deionised water to remove any foreign organic matter, and then chopped into small pieces in a petri dish. Pneumatophores were also thawed and rinsed with deionised water. Epiphytes were scraped from the surface taking care not to remove pneumatophore bark, and rinsed and retained in a petri dish. Final effluent samples were filtered through pre-ashed glass fibre filters (GF/C, 450oC for 4hrs) with the filtrate retained in an aluminium oven dish, and each filter (with filtered residue) placed into a glass petri dish. All samples were dried at 65oC for 48 hours, or until fully dried. Filters were placed whole into a labelled glass vial, otherwise all remaining dried samples were ground to a fine powder on a Retsch MM200 mixing mill. For difficult samples, where small amounts of sticky residue remained, grinding was done manually using mortar and pestle.

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Sub-samples of each dried mangrove and epiphyte sample were weighed into a tin capsule (7.00±1.00 mg) then tightly sealed and placed into a 96-well plastic tray. All samples were sent to Griffith University Stable Isotope Laboratory (Nathan Campus, Qld) for δ15N and δ13C stable isotope analysis using a Continuous Flow Infra Red Mass Spectrometer (CF-IRMS). Data are reported relative to International Atomic Energy Agency secondary standards calibrated against global standards of Vienna PeeDee Belemnite for carbon, and atmospheric air for nitrogen. Stable isotope values are reported in delta (δ) units, in parts per thousand (‰) relative to the international standard and determined as per the equation:

= × 1000 1 𝑅𝑅𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠 𝛿𝛿𝑋𝑋 � � 𝑅𝑅𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠𝑠 − Where X is carbon or nitrogen and R is the ratio of the heavy isotope over the light isotope. Results are accurate to 0.1% for both C % and N% and ± 0.3 ‰ for δ13C and δ15N.\

DATA ANALYSIS

Water quality parameters by site were compared to distance upstream from creek mouth, and any observed relationships tested for correlation using JMP 13.0.0 (SAS Corporation 2016). Where linear relationships were observed, regression analysis in Prism 7 (Graphpad Software, 2017) was used to quantify the relationship. Principal Component Analysis was conducted on all water quality parameters to observe the main interactions between parameters. Results were compared to historical water quality data from Lake Macquarie City Council (LMCC), for a lone monitoring point on Cockle Creek (see LMCC on Figure 5), and compared to National Water Quality guidelines (ANZECC & ARMCANZ 2000) for recreational and aesthetic purposes.

A paired t-test was conducted using JMP (SAS Corporation 2016) on mean site mangrove leaf δ15N values for both the pilot and August sampling. As this test confirmed a shift in isotope values between February and August (Appendix A), pilot data was not included in subsequent analysis. The interpolated map produced for pilot δ15N is shown at Appendix A. Data for the second passive mangrove leave collection were tested for normality, and δ15N transformed using the natural logarithm function (ln), prior to ANOVA testing in JMP 13.0.0 (SAS Corporation 2016) to determine the statistical significance of observed differences in δ15N and δ13C by site.

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Passive epiphyte data was compared to passive mangrove δ15N results and tested for correlation. Isotope data for these epiphytes was tested for normality, prior to ANOVA testing in JMP 13.0.0 (SAS Corporation 2016), to determine the statistical significance of observed differences in δ15N and δ13C by site.

For deployments of epiphytes as active bio-indicators (May & June), results were tested for normality, and ANOVA testing conducted as above for passive epiphyte data. Wilcoxon/Kruskal-Wallis Tests for non-parametric data were used to determine any differences between sites for the active deployment in June. Mean δ15N and δ13C values for each deployment were compared, and the mean δ15N results were further contrasted against rainfall data. Further, mean δ15N and δ13C values obtained from the reference site were subtracted from raw data for each sample to obtain an absolute enrichment value per sample. This value was used to calculate mean enrichment per site, which was compared to site distance upstream.

The ratios of carbon to nitrogen (C:N) of all samples were calculated from elemental data (by weight) to monitor for changes to nutrient source. Mean C:N for each site were compared to distance from creek mouth, and any differences between sites assessed for significance with ANOVA testing in JMP 13.0.0 (SAS Corporation 2016). Variations by site for isotope data (δ15N and δ13C) were compared to variations in C:N, examining for similar patterns by site. A bi-plot graph of δ15N vs δ13C was created, and relationships between similar sites on this plot were examined.

SPATIAL ANALYSIS

The upstream distance of sites from the mouth of Cockle Creek were determined using the measuring tool in QGIS 2.10.1 (www.qgis.org). Site 15 opens to Cockle Bay on Lake Macquarie, and was set as distance 0m. Site locations were imported into QGIS and individual parameters (water quality, stable isotope data) displayed in the main layer. Data was then exported to ArcGIS (ESRI 2016), and interpolated using the Inverse Distance Weighting (IDW) tool. A barrier line feature layer was added to prevent values from the weir interfering with interpolation in the main channel. The resulting raster images were clipped to the Cockle Creek / Lake Macquarie outlines, and displayed in maps using continuous colour stretch. A plume map was produced for mean δ15N in passive mangrove leaves (August),

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passive epiphytes (February), and mean change in δ15N for each active epiphyte deployment (May and June).

RESULTS

PASSIVE BIOINDICATORS

Isotope data in mangrove leaves

δ15N of mangrove leaves varied between 3.8 ‰ and 13.3 ‰ (sites J, N; Figure 8). There were significant differences among sites after log transformation (F = 13.084, p < 0.001), with 5 sites ( i, G, J, 6 and 10) significantly higher than all other sites. In the central part of the creek, δ15N is elevated at the WWTP (site 6), decreases sharply at the next site downstream before steadily increasing to a maximum over 1500m downstream.

Figure 8: Mean (± SE, n=3) δ15N in mangrove leaves by site against distance upstream. The mean value of recycled water from Edgeworth WWTP is shown as a grey line (WWTP, 9.5 ‰), while the grey shaded area represents the range of reported mean δ15N values in mangrove leaves for natural or urban catchments (Fry & Cormier 2011, Kousbrouek et al. 2017, Reis et al. 2017).

Values for δ13C ranged between -31.7 and -26.9 ‰. A weak but significant trend of decreasing δ13C moving upstream was observed (R2 = 0.209, F = 5.80, p = 0.025; Figure 9). When comparing δ15N vs δ13C in mangrove leaves, sites 14, G, 6, I and J had slightly higher

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δ15N and δ13C than other sites and were distinguishable from the other sites. Site C had lower values of both than other sites, and was also distinguishable (Figure 10).

Figure 9: Mean δ13C (± SE, n=3) in mangrove leaves by site against distance upstream. A weak but significant trend of decreasing δ13C moving upstream is observed ( r2 = 0.209, F = 5.802, p = 0.025), and modelled by the grey line.

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Figure 10: For mangrove leaves as passive bio-indicators, mean (± SE, n=3) δ15N by site against mean (± SE, n=3) δ13C by site. δ15N values varied more widely in comparison to the much narrower range of δ13C.

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Isotope data in passive epiphytes

δ15N in passive epiphytes varied between 2.0 ‰ and 7.5 ‰. Sites 5 and 10 were significantly higher than all others, while sites 12, 14 and 15 were significantly lower than all others (F = 2.948, p = 0.005; Figure 11). Passive epiphyte δ15N had a weak correlation to mangrove leaf δ15N (R2 = 0.186, F= 6.146, p= 0.020). There is a weak trend of increasing δ15N with distance upstream for passive epiphytes in February (R2 = 0.165; p = 0.005 , Figure 11).

δ13C in passive epiphytes (February) ranged between -32.0 ‰ and -22.6 ‰. Differences in δ13C between sites was significant (F= 2.061, p= 0.044), but showed no spatial pattern..

8.0

7.0

y = 0.0002x + 3.837 6.0 R² = 0.165

N‰ 5.0 5 1

δ

4.0 Epiphytes (passive)

3.0

2.0 0 1000 2000 3000 4000 5000 6000 7000 8000 Distance upstream (m)

Figure 11: Mean δ15N (± SE, n=3) in passive epiphytes collected from banks of Cockle Creek against distance upstream. A weak trend is observed.

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SPATIAL RESULTS IN PASSIVE BIOINDICATORS

Mangrove leaves

Spatial interpolation shows hot spots of elevated δ15N close to Edgeworth WWTP, and separately, another hotspot along the central creek body. A further hotspot, not as intense, is observed downstream (Figure 12).

Figure 12: Interpolated δ15N values in mangrove leaves in Cockle Creek. Biological influence around the WWTP is localized (300m range), and an elevated δ15N source appears downstream.

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Epiphytes (Passive)

Passive epiphytes showed similar distribution of δ15N (Figure 13) as for mangrove leaves, but with only one hotspot in the main creek body.

Figure 13: Interpolated δ15N in passive epiphytes collected from the creek banks, February 2017, showing similar distribution patterns to mangrove leaves, but only one hotspot in the main body of the creek.

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ACTIVE BIO-INDICATORS

Reference data for active deployment of epiphytes

Baseline values for δ13C and δ15N in epiphytes obtained at the reference site (Dora Creek) were similar on both dates, however, C:N showed some variation between dates (Table 2).

Table 2 : Summary of mean results and Standard Error of the Mean (SEM) for epiphytes from reference site. Mean δ13C SEM Mean δ15N SEM Mean C:N SEM (n=5) (n=5) (n=5) May (dry) -28.7 0.5 0.2 0.2 29.8 2.9 June (wet) -28.0 0.8 0.1 0.3 24.1 2.3

Isotope data for active bioindicators

For active epiphyte deployments, δ15N values were much lower than passive epiphytes. In May (dry weather), δ15N ranged between -0.5 - 1.9 ‰ and in June (wet) from -1.2 to 2.4 ‰ (Figure 14). There were no significant differences between sites in May (F= 0.923, p= 0.54) or June (Wilcoxon/Kruskal-Wallis test for non-parametric data; 2 = 17.944, p = 0.13), or between the two sampling dates (t = 0.207, p = 0.418). During activeχ deployment, all sites showed an increase in δ15N relative to the reference site value, except site 1 and site 11 in June.

For the active deployments, δ13C values ranged between -30.6 ‰ and -25.2 ‰ in May, and from -31.2 ‰ to -26.7 ‰ in June. No apparent trends or differences from the reference values (-28.7 ‰ in May, -28.0 ‰ in June) were observed. Differences in δ13C between sites was significant for June (F = 3.253, p= 0.006), but not May. There was no difference in mean δ13C between the two active deployments in May and June.

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2.5 A May deployment 2.0 June deployment 1.5

1.0 N‰

5 0.5 1

δ 0.0

-0.5

-1.0

-1.5 0 1000 2000 3000 4000 5000 6000 7000 8000 Distance upstream (m)

2.5 May deployment B 2.0 June deployment 1.5

1.0 N‰ 15

δ 0.5

0.0

Change in Change -0.5

-1.0

-1.5 0 1000 2000 3000 4000 5000 6000 7000 8000 Distance upstream (m) Figure 14: (A) Mean (± SE, n=3) raw δ15N in epiphytes for each active deployment in May and June against distance upstream; (B) Mean (± SE, n=3) change in δ15N in deployed epiphytes, relative to average reference value obtained at Dora Creek for that deployment date.

SPATIAL RESULTS FOR ACTIVE BIOINDICATORS

Individual active deployments in May and June show generally similar spatial patterns for interpolated change in δ15N, but after wet weather in June, effects were more widespread (Figure 15). The distributions are similar to those for passive indicators (Figure 12 and 13) above.

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Figure 15: Change in δ15N(‰) in active bio-indicator epiphytes, for May (top) and June 2017 (bottom).

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C:N RATIOS

Mangrove leaves

Mean C:N ratios in mangrove leaves varied between 16.8 (site 14 near the mouth) and 34.9 (site E), however, there were no significant differences between sites (F = 1.68, p = 0.064) or with distance upstream (Figure 16).

Figure 16: Mean (± SE, n=3) ratio of C:N in mangrove leaves by site against distance upstream.

Epiphytes (passive)

For the passive epiphyte collection, mean C:N ratios varied greatly between 9.8 (site 6 near the WWTP) and 45.5 (site 10), with no significant patterns. Mean C:N in passive sites across the estuary was 20.7.

Active bioindicators

For the active deployments, mean C:N was 24.9 and 22.1 for May and June respectively, and differences between deployments were significant (t = -2.372, p = 0.023 ; Figure 17).

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40 May deployment

35 June deployment

30

25

20 C:N ratio 15

10

5

0 0 1000 2000 3000 4000 5000 6000 7000 8000 Distance upstream (m) Figure 17: Mean (± SE, n=3) C:N ratios in active epiphyte deployments by site against distance upstream.

WATER QUALITY

Measurements for pH were not available for May and June sampling. In February, a strong trend of increasing pH towards the creek mouth was present (R2 = 0.71 , p < 0.01), but not in August. All values were within ANZECC guidelines of 6.5 – 8.5. In August pH decreased moving downstream from site A (in the vicinity of the WWTP) dropping to a minimum value recorded at Site 10 (2109m from the creek mouth). At site i, 116m downstream from site 10, pH rose to 7.37, then increased steadily towards the creek mouth (Figure 18).

Salinity was lowest (28.39 ppt) at Site 1, the farthest site upstream at the weir, on 30 June after wet weather. It was highest (35.65 ppt) at site 14 near the creek mouth, on 30 August. Site 1 typically had the lowest recordings for salinity on any given sampling date. Differences between sampling dates were observed, particularly for June sampling dates after rain, where salinity was generally lower and more variable. Other dates showed a typical estuarine trend of reducing salinity moving upstream (significant negative relationship, R2= 0.54, P < 0.001, Figure 19).

Temperature varied on a seasonal basis, with water temperature in February reaching 30.1°C at site 1, but decreasing to 14.3 °C in June. For each sampling date, temperature decreased moving downstream, however, site 1 (furthest upstream) often defied this trend., being colder relative to other sites in May and June, and most notably after heavier rain in June (Figure 21).

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8.5

1 Feb 8 30 Aug

7.5 pH (pH (pH units) pH

7

6.5

6 0 1000 2000 3000 4000 5000 6000 7000 8000 Distance upstream (m)

Figure 18: Measured pH vs distance upstream at examined sites, at different sampling dates. with a strong trend of increasing pH downstream for February (R2 = 0.71 , p < 0.01). A depression in pH can be seen in the central estuary (2000 – 4000m upstream) on 30 August.

Conversely, dissolved oxygen decreased moving upstream, with some variability occurring mainly in the middle estuary (Figure 21). Measurements ranged from 20.6% (site 1, 11 May) to 122% (site 12, 11 May). Most readings were above ANZECC guidelines of 80%, except at the central creek junction adjacent to the WWTP, and the upper arm of Cockle Creek, notably at the weir, where readings were consistently lower.

Chlorophyll a and turbidity were variable across the estuary. Chlorophyll ranged from 0.03 to 1.24 μg/L (ANZECC trigger is 5 μg/L), showing no obvious trends, while turbidity ranged from 2.9 - 57.0 NTU, with ANZECC trigger values of 1-100 NTU.

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38

1-Feb

36

5 May

34 11 May

32 23 Jun

30 30 Jun Salinity per thousand) (parts Salinity

30 Aug 28

26 0 1000 2000 3000 4000 5000 6000 7000 8000

Distance upstream (m)

Figure 20: Measured salinity by date against distance upstream at examined sites.

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1 Feb

30 5 May

11 May

)

C ° 25 23 Jun

30 Jun

30 Aug

Temperature ( 20

15

10 0 1000 2000 3000 4000 5000 6000 7000 8000

Distance upstream (m)

140 1 Feb

120 5 May

11 May

100 23 Jun

80 30 Jun

30 Aug 60 Dissolved oxygen Dissolved oxygen %

40

20

0 0 1000 2000 3000 4000 5000 6000 7000 8000 Distance upstream (m)

Figure 21: Measured temperature by date against distance upstream at examined sites (TOP) generally increased moving upstream, while (BOTTOM) measured dissolved oxygen vs distance upstream at examined sites decreased.

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DISCUSSION

The use of active and passive stable isotope bio-indicators revealed the extent of anthropogenic inputs in Cockle Creek. Specifically, the use of recycled water onsite at Edgeworth WWTP and at Waratah Golf Club is not contributing to the observed elevated 15N in Cockle Bay, and the influence of alternate sources within the catchment became clear under spatial analysis. The zone of influence of sewage-derived nutrients from Edgeworth WWTP was found localised to a range of 300m, which may increase slightly following rainfall. Interpolated maps were presented showing the extent of this zone, as well as the presence of other sewage derived nutrient sources. Amongst other industrial sources in the estuary, it is likely that drainage from the Pasminco rehabilitation site, where recycled water is stored and used on site, is impacting on δ15N values in the main body of Cockle Creek.

Mangrove leaves proved effective and reliable as passive bio-indicators, offering a view of average ambient nutrient conditions for the last 12-18 months. They allowed quantitative and spatial assessment of 15N enrichment in Cockle Creek. Epiphytes growing on mangrove pneumatophores were also effective as passive bio-indicators, showing areas of relative enrichment, and allowing spatial assessment that agreed closely with mangrove leaf results. Some differences were noted owing to the much shorter temporal resolution of epiphyte tissue.

Active deployment of epiphytes requires longer than the 5 days afforded in this study, in order to allow full acclimation with ambient conditions. A suggested deployment time of 14 days is recommended in future study. Notwithstanding, epiphyte deployments allowed spatial assessment of 15N enrichment relative to reference values, which agreed closely with passive mangrove leaves and passive epiphyte bio-indicators.

The significance of this study is introducing and refining novel and effective methods to detect sewage-derived nutrient sources, and delineate their zone of biological influence. At the same time, it provides much-needed information on the impacts of recycled water application near an estuary. Both of these outcomes will be of great value for estuary users and managers, WWTP operators and recycled water users.

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PASSIVE BIO-INDICATORS

Mangrove leaves proved effective as passive bio-indicators. Variation between site replicates was generally low, confirming the suitability of mangroves as reliable indicators of nutrient history (Fry & Cormier 2011). Passive epiphytes harvested in situ also provide an accurate spatial picture of relative ambient nutrient source history over a short time frame – days to weeks - and physicochemical parameters. Both methods were able to detect significant differences between sites in 15N levels. Neither method was able to resolve any change in carbon sources.

In a spatial sense, Edgeworth WWTP at the main creek junction was a hotspot for leaf δ15N, which dissipated within 300m in all directions. Sites in this region had a δ15N signature matching very closely with the WWTP output, which is likely the dominant nutrient source in this vicinity. Downstream of the railway bridge, adjacent to a LMCC materials depot (Figure 5, “Brownfields”), was a hotspot with much higher δ15N, extending downstream for several hundred metres. A number of road-drains run into this section, and part of the Pasminco rehabilitation site also drains, via retention ponds, through this area (Figure 5). A third major hotspot was noted further downstream, over a similar length of creek facing the Council maintenance depot at Boolaroo, albeit with lower δ15N signals. Several light industries are located nearby, as well as a small concrete drain from Boolaroo and Speers Point, which is subject to tidal inundation.

Immediately downstream of this elevated δ15N hotspot, sites showed relatively low δ15N, as did sites on the marsh islands in the creek mouth delta. This pattern was seen also in the pilot study. While not low compared to results reported by other authors, the relative values in relation to nearby hotspots would indicate some other source input of lower δ15N levels. Very high levels of ammonium (Fry & Cormier 2011), dominant in mangrove marsh pore water (Fry et al. 2000), is linked to high fractionation, leading to reduced plant (or algal) δ15N. This seems likely for sites on the delta marsh island.

Raw sewage effluent or excretion could also be a source of relatively low δ15N (Fry 2006). Hunter Water infrastructure carries treated effluent or recycled water with increased δ15N, and can be eliminated as a source, but leaking or illegal sewer connections from residential or

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commercial properties, diverting raw sewage into the stormwater system, are a possible alternate source.

Results for δ15N in passive epiphytes showed very similar spatial distribution to A. marina leaves, and correlated δ15N values, but epiphyte δ15N was not as high as leaf δ15N. This is in contrast to previous authors (Daigle et al. 1990, Bouillon et al. 2008). The third or lower hotspot observed in leaf results was not observed at all in passive epiphyte results. Some differences are to be expected due to species-specific fractionation processes, but more importantly, the different timeframes for nutrient assimilation between leaves and epiphytes will be resolving different time periods, and potentially indicate changes to ambient nutrient conditions.

The map of leaf data in August (Figure 16) was consistent with that from the pilot study in February (Appendix A) but with significantly higher δ15N results. High rainfall events were experienced in March and June of 2017 (Figure 7), which could have driven observed changes to δ15N. This is discussed below. Although significant, the increase between δ15N in February’s pilot and August was, at most sites, limited in scale to less than 1‰. This increase could be due to effects of rain, or seasonal changes in A. marina nutrient pathways (Macko et al. 1987, Yamaguchi et al. 2017), although Kousbrouek et al. (2017) found no significant temporal differences in A. marina δ15N.

The maximum mangrove leaf δ15N value of 13.3 ‰ was double the maximum obtained by Kousbrouek et al. (2017) in Tilligerry Creek, a NSW estuary also receiving WWTP-derived nutrient inputs. Although many sites, both in August and the February pilot were in a similar range as at Tilligerry, some sites were much higher. Reis et al. (2017) reported mean δ15N in leaves between 2.7 and 4.3 ‰ (SD 1.9 – 2.1). In Hawaii, Fry and Cormier (2011) found δ15N of 5.9‰ in an urban catchment with red mangrove leaves, and 11.6‰ in a golf course impacted catchment, comparable with present results.

Reis et al. (2017) reported mean C:N ratios from 26.3 to 30.9 (SD 4.6 – 5.8) in mangal forests in Brazil, while Cifuentes et al. (1996) reported C:N ratios for mangroves in Ecuador between 40 and 75. Present study results fall between these two ranges

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δ13C in mangrove leaves showed some variation, but not enough to discern any pattern. Only site C showed any real variation from the mean δ13C, perhaps a result of the low flat marshland adjacent to the creek at this site providing alternate carbon dynamics (Signa et al. 2017). Otherwise, strong photosynthetic fractionation appears to regulate leaf δ13C for A. marina in Cockle Creek (Mariotti et al. 1984, Bouillon et al. 2008).

ACTIVE BIO-INDICATORS

As active bio-indicators, mangrove pneumatophore epiphytes were moderately successful. Enrichment responses in deployed epiphytes followed similar patterns to passive bio- indicators. Observed differences in enrichment responses were not significant, most likely due to insufficient deployment time for acclimation. In a spatial sense, however, the locations of relative hot spots did align with those from the passive collection, where differences were significant. Specific locations were similar but not identical to passive bio-indicator maps, but they appeared to show an accurate short-term picture of δ15N distribution, in contrast to the longer temporal resolution showed by mangrove leaves.

Also, spatial distributions for each of the two active deployments showed similar characteristics, but with apparent differences. Specific locations were similar but not identical. The two hotspots in the lower main channel for both deployments were consistent with passive results, as were relatively depleted sites. In contrast, the vicinity of the creek junction and golf course, a hotspot in passive data, showed virtually no increases in δ15N in May, but was more elevated in June after rainfall. In May, a hotspot directly south of the WWTP on upper Cockle Creek was replaced in June by a larger, more intense, hotspot approximately 1km upstream. This could be the result of an undetermined change to catchment inputs, or tidal action moving WWTP nutrients upstream.

There is a drawback to using epiphytes as indicators. Changes to nutrient sources can alter species assemblages (Boon & Bunn 1994, Melville & Connolly 2003), introducing uncertainty regarding fractionation rates and tissue turnover rates at sites where alternate source inputs are suspected. Further work to quantify the change to fractionation rates or turnover times under changing species composition would allow more quantitative conclusions from actively-deployed epiphyte data.

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Laboratory determined turnover times – being species composition dependent - are likely to vary geographically and temporally. The deployment of algal traps can assist by controlling the ambient nutrient timeframe. However, although five days was adequate in Queensland’s spring and summer(Costanzo et al. 2001), it was too short for complete acclimation in this study, limiting the ability to make quantitative conclusions.

Although based on absolute increase in δ15N, uncertainty regarding acclimation means no conclusions can be made regarding the absolute δ15N value of suspected sources. Caution must be taken in interpreting these results, as more rapid enrichment seen at some sites could be due to large volumes of slightly elevated source nutrients, or low volumes of extremely elevated nutrients. With the limitations of the short deployment times, conclusions can only be made on the relative spatial differences in absolute enrichment (from the reference standard) between sites.

Further study, based on 14-day deployment as used by Rush (2003), would allow full acclimation to ambient nutrient conditions, and more quantitative conclusions about dominant sources. One drawback to increased deployment time, though, is the increased likelihood of losses. Two traps were lost during this study (site 14 in May, and site 7 in June), with no sign of the apparatus or weights. Both were in locations with relatively easy shore access.

Little specific literature on mangrove epiphyte communities and associated fractionation processes exists. This potentially limits the scope of conclusions, however epiphyte δ15N in this study was comparable to a range of studies involving algal or other epiphyte communities (Daigle et al. 1990, Cifuentes et al. 1996, Fry et al. 2000, Costanzo et al. 2001, Divers et al. 2014).

POTENTIAL INPUTS AND SOURCES

Edgeworth WWTP has a clear, yet localised biological influence on Cockle Creek. The range of this effect appears to extend approximately 300m in all directions, where δ15N levels are indistinguishable from background levels in the estuary. Rainfall may lead to temporal changes in this nutrient distribution, but these changes do not appear to have significant impact on the long-term ambient nutrient history.

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Excess recycled water irrigation at Waratah Golf Course is not a significant nutrient source in the estuary. For almost the entire 2.5km frontage in Cockle, Brush and Winding Creeks, δ15N levels are comparatively low. Observations of the course during field sampling noted the course is not particularly lush, nor green. Evidence and signs of algal blooms in the creek were noted in places, but δ15N results show this is not related to recycled water use or effluent discharge.

At the southern end of the golf course, δ15N levels begin to increase, apparently independent of the WWTP. The creek joining Cockle Creek at this site runs through the course, but has its origins at the Pasminco rehabilitation site, a large site, with several distinct drainage paths, some sub-surface due to past industrial filling practices, leading to the Cockle Creek system (Ryall 2005). A milky, opaque plume was observed emanating from the drainage creek on 5 May, and on 23 June after wet weather, the plume extended downstream for several hundred metres, keeping in a visibly distinct channel within five metres of the eastern shore of Cockle Creek.

Work has been ongoing at the Pasminco site for several years now, with significant loads of contaminated top soils moved into a containment cell, and the creation and relocation of interim retention dams and drainage works. Restoration also involves cover with new top soils, and associated re-vegetation (Manidis Roberts 2010). Given the industrial history and contamination of the site, effects of those activities on isotope levels are difficult to predict. Fertilizer production, which occurred on-site for several decades (Manidis Roberts 2010), normally fixes atmospheric nitrogen, and thus would generally lead to a reduced δ15N signature (Fry 2006) if it entered Cockle Creek, or was being applied to the catchment. Fertilizer production is therefore not responsible for any observed enrichment.

Recycled water is an integral component of the Water Quality and Water Cycle Management Plan for the Pasminco site (Murphy 2006). Further, it is stored on site in dams, exposing it to highly selective evaporation, which would increase δ15N further. Additionally, an Effluent Treatment Plant (ETP) on-site processes 30L per hour of contaminated dam water, mixing it with pumped water from Cockle Creek before disposing of output effluent into the creek via EPA-licensed discharge (Murphy 2006). Consistent with findings of elevated δ15N at all

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places where the restoration site is connected to the creek, via drainage, it can be concluded that recycled water applied for sediment control and revegetation is finding ways into Cockle Creek, resulting in elevated δ15N observed in mangroves and epiphytes in the creek.

Manidis Roberts (2010) identified a risk of groundwater interactions from the Pasminco site, but due to ongoing water extraction at the time for coal handling, there was no connectivity between the creek and the lowered groundwater. They noted the risk must be revisited if, as has occurred, water extraction ceased. The possibility of groundwater interactions cannot be ruled out altogether, although diffuse sources would tend to elevate δ15N throughout the estuary.

The coal handling plant is also adjacent to the main body of Cockle Creek, and may be an important contributor to δ15N variations. Although coal handling at the washery ceased in 2016, the waters in the dams surrounding it were known to have a high content of recycled water supplied by Hunter Water (Hunter Water 2011). During high rainfall, these dams can overflow into the main creek (observed twice during June), however, this is not evident in δ15N levels. Adjacent to the overflow δ15N was consistently near the average readings for passive leaves and epiphytes, although slightly elevated on active epiphyte deployments. When active, the washery may have been a contributing source of sewage-derived nutrients, but this is not apparent near the overflow point since its closure. This would imply leaf turnover has fully cycled since operations ended in May 2016, which is at the lower end of the expected time (Ellison 2002, Alongi et al. 2005a, Kumar et al. 2011). Alternatively, nutrients have been captured in sediments in the dams, only to be released in extreme storm events.

Another possibly unrelated source appears to be impacting further downstream. Although not as elevated as either of the other major hotspots, sites at the lower end of the creek before the main channel splits showed consistently high δ15N levels. One site is adjacent to a large underground suburban drain entering the creek, and elevated δ15N could be the result of some source from within the urban drain catchment, which includes part of the nearby industrial estate and council works depot, and parts of Boolaroo. It may also be connected to the Pasminco site by drainage. Another site is located adjacent to a small, open cement drain extending several hundred metres into Boolaroo and Speers Point. The drain experiences some tidal inundation, and has partially submerged deposits of wet muds and sediment for up

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to 50m along its length. Shallow water is subject to increased evaporation, which is highly selective, and increased sediment interactions (nitrification and denitrification), which can also have an effect on fractionation rates. It is possible that the physical characteristics of the drain might contribute to observed increases in δ15N.

The activities of waterway users and residents may also be an important factor in observed δ15N values. Unauthorised drainage and bank construction is common, and lawn clippings and green waste deposited on banks were a regular sight when sampling. The upper navigable section of Winding Creek in particular was noted for thick green grass to the water’s edge, in stark contrast to the golf course across the creek. In this area, δ15N was generally low, consistent with fertiliser use, high levels of ammonium, or raw effluent being diverted into the creek.

EFFECT OF RAINFALL/RUN-OFF

Whatever the impact of rainfall on sewage-based nutrient distribution, it has a limited effect on the observed long-term nutrient history. A small increase was recorded in leaf δ15N between the February pilot and August sampling, for which seasonal influences cannot be discounted. If rainfall and associated catchment run-off is the cause of a general increase in δ15N values, then it is reasonable to expect they will gradually decrease over time without rain. Indeed, the nutrient history obtained in the pilot study included massive rainfall in January 2016 (Appendix B), and further heavy falls in June that year, which would suggest general pulse increases to δ15N with heavy rainfall, followed by a gradual decrease between rainfall events.

If, during very large storm events, surplus nitrogen from storm events is no longer rate limiting, increased variability in δ15N, due to preferential nitrogen uptake mechanisms (Rush 2003), or through nitrification and de-nitrification (Mariotti et al. 1984), could also be expected. Although this may confound quantitative interpretation of rainfall results, it could be reasonably expected that much of the surplus nitrogen would be carried out of the estuary with flood flow, and conditions would generally be restored over time without rain, as above. Figure 7 shows that after rain that flooded the weir before the June deployment, tidal variation at Cockle Creek Station was quickly restored. This would imply that normal estuarine

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processes are restored, and that excess transient water column nutrients (rainfall driven) have quickly settled or been dispersed.

For some sites, however, δ15N was consistently elevated independent of rainfall. This implies a constant source of elevated δ15N within the catchment; either through leaking sewage effluent assets (probably the case around the creek junction), or through irrigation of δ15N enriched products within the catchment. This agrees with findings by Clarke (2016), that in contrast to other sites within Lake Macquarie, where rainfall events generally triggered increased δ15N, Speers Point actually decreased, indicating dilution of the regular elevated signature by rain from the catchment. The mean δ15N for site J of 13.3 is much higher than that observed for the WWTP effluent, further implying an alternate source within the catchment.

This is particularly evident in the main channel. Hotspots adjacent to the Pasminco site, and the industrial area downstream, showed consistently elevated δ15N signatures, relative to other sites within the estuary, the WWTP effluent signature, and also previously reported A. marina results.

WATER QUALITY PARAMETERS

Water quality data were variable, but with expected estuarine trends in temperature, pH, dissolved oxygen and salinity evident with increasing distance from the creek mouth. Similarly, correlation in chlorophyll a and turbidity was to be expected (Scanes et al. 2007).

The results for pH are similar to results from the Lake Macquarie City Council (LMCC) monitoring program (Appendix B) where pH varied between 7.8 and 8.4. With no recent published data beyond the road bridge, no explanation is apparent for the depressed pH in the main channel of Cockle Creek on August 30, but rainfall driven drainage is not responsible, given the dry period before the sampling. Given the consistent elevated δ15N signal in that segment of the creek, it may indicate an alternative, irrigation-based input to the estuary, acting locally downstream of the railroad bridge crossing. Alternatively, with pH potentially influencing fractionation rates (notably during photosynthesis), this has the potential to confound results, if pH were consistently depressed. This was not the case on other sampling dates.

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Any effect due to rainfall (in June), or prolonged dry (in May), on water quality parameters (perhaps other than salinity) could not be discerned except at the Barnsley Weir. Chlorophyll a and turbidity both being markedly higher in summer would appear to be temperature controlled, and again is consistent with patterns in LMCC data.

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CONCLUSION

There are several apparent sources of variation in δ15N in Cockle Creek. The associated uncertainty with sources, and their δ15N signal, prevents quantitative conclusions being drawn, although the localised area of biological influence for the WWTP (approximately 300m in all directions) is unequivocal. Heavy rainfall events may expand this area, but levels reduce again without rain, and the effect on long-term δ15N is muted. It would also seem clear that recycled water in use on the golf course has no detectable biological effect on the estuary. Occasional blooms may occur, but residential and waterfront activities in this part of the creek may be more significant factors.

Other sources within the catchment, most noticeably the Pasminco rehabilitation site, appear to be linked to elevated δ15N signals in the main body of Cockle Creek. Further urgent work is required to review the cessation of water extraction, subsequent to the termination of coal handling operations, and determine the impact on groundwater levels. Groundwater interactions could have the potential to introduce more complex sources, and confound any conclusions drawn in this study. More concerning, is the known susceptibility of the Pasminco site to groundwater interactions, and the potential for other contaminants to leak.

Mangroves proved a reliable, effective and readily sampled bio-indicator for stable isotope analysis, suited to provide useful long-term history of ambient nutrient sources in the medium term of 13 months or more. Epiphytes associated with mangrove pneumatophores, a novel method, also proved effective at determining relative enrichment in a spatial sense. Further work is required to determine exact tissue turnover and fractionation rates for particular species composition if quantified conclusions are required. Notwithstanding, for a recent ambient nutrient source history, both harvested passively in situ and actively deployed epiphytes were reliable indicators of areas of relatively higher δ15N. A suggested deployment time of 14 days is recommended in NSW estuarine waters.

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REFERENCES

Alongi DM, Clough BF, Robertson AI (2005a) Nutrient-use efficiency in arid-zone forests of the mangroves Rhizophora stylosa and Avicennia marina. Aquatic Botany 82:121-131 Alongi DM, Ramanathan AL, Kannan L, Tirendi F, Trott LA, Bala Krishna Prasad M (2005b) Influence of human-induced disturbance on benthic microbial metabolism in the Pichavaram mangroves, Vellar–Coleroon estuarine complex, India. Marine Biology 147:1033-1044 ANZECC, ARMCANZ (2000) Australian and New Zealand guidelines for fresh and marine water quality. In: Australia E (ed). Environment Australia, Canberra AWACS (1995) Lake Macquarie Estuary Process Study. In: Living Lake Macquarie. Lake Macquarie City Council, Speers Point, NSW Azad MAS, Ancev T (2010) Using ecological indices to measure economic and environmental performance of irrigated agriculture. Ecological Economics 69:1731-1739 Boon PI, Bunn SE (1994) Variations in the stable isotope composition of aquatic plants and their implications for food web analysis. Aquatic Botany 48:99-108 Bouillon S, Connolly RM, Lee SY (2008) Organic matter exchange and cycling in mangrove ecosystems: Recent insights from stable isotope studies. Journal of Sea Research 59:44-58 Carpenter SR, Caraco NF, Correll DL, Howarth RW, Sharpley AN, Smith VH (1998) Nonpoint pollution of surface waters with phosphorous and nitrogen. Ecological Applications 8:559-568 Cifuentes LA, Coffin RB, Solorzano L, Cardenas W, Espinoza J, Twilley RR (1996) Isotopic and Elemental Variations of Carbon and Nitrogen in a Mangrove Estuary. Estuarine, Coastal and Shelf Science 43:781-800 Clarke A (2016) Seagrass in Lake Macquarie. PhD thesis, University of Newcastle, Ourimbah, NSW Clibborn B (2017) OCAL Complex Annual Review 2016. Teralba Costanzo SD, O’Donohue MJ, Dennison WC, Loneragan NR, Thomas M (2001) A New Approach for Detecting and Mapping Sewage Impacts. Marine Pollution Bulletin 42:149-156 Costanzo SD, Udy J, Longstaff B, Jones A (2005) Using nitrogen stable isotope ratios (δ15N) of macroalgae to determine the effectiveness of sewage upgrades: changes in the extent of sewage plumes over four years in Moreton Bay, Australia. Marine Pollution Bulletin 51:212-217 Creighton C, Boon PI, Brookes JD, Sheaves M (2015) Repairing Australia’s estuaries for improved fisheries production – what benefits, at what cost? Marine & Freshwater Research 66:493-507 Daigle ST, Fleeger JW, Cowan JH, Pacsal P (1990) The epiphyte community of mangrove roots in a tropical estuary: distribution and biomass. Aquatic Botany (Netherlands):117 Day JH (1980) What is an estuary? South African Journal of Science 76, 198.

Recycled water in Cockle Creek Page 52

DEC (2004) Use of Effluent by Irrigation. Department of Environment and Conservation, Sydney, NSW Deeley DM, Paling EI (1999) Assessing the ecological health of estuaries in Australia. National River Health Program. Marine and Frshwater Research Laboratory, Murdoch University, Perth Divers MT, Elliott EM, Bain DJ (2014) Quantification of Nitrate Sources to an Urban Stream Using Dual Nitrate Isotopes. Environmental Science & Technology 48:10580-10587 Ellison AM (2002) Macroecology of mangroves: large-scale patterns and processes in tropical coastal forests. Trees 16:181-194 EPA (2017) EWWTP EPA License 2017. ERM Australia (2000) Five Islands Road Project Environmental Impact Statement. Book 1. Roads & Traffic Authority, Sydney, NSW Fry B (2006) Stable Isotope Ecology. Springer, New York, p 40-270 Fry B, Bern AL, Ross MS, Meeder JF (2000) δ15N Studies of Nitrogen Use by the Red Mangrove, Rhizophora mangle L. in South Florida. Estuarine, Coastal and Shelf Science 50:291-296 Fry B, Cormier N (2011) Chemical Ecology of Red Mangroves, Rhizophora mangle, in the Hawaiian Islands. Pacific Science 65:219-234 Gaston TF, Kostoglidis A, Suthers IM (2004) The 13C, 15N and 34S signatures of a rocky reef planktivorous fish indicate different coastal discharges of sewage. Marine and Freshwater Research 55:689-699 Guest MA, Connolly RM, Loneragan NR (2004) Short communication: Within and among-site variability in δ13C and δ15N for three estuarine producers, Sporobolus virginicus, Zostera capricorni, and epiphytes of Z. capricorni. Aquatic Botany 79:87-94 Hodgson BR (2003) Nearshore sediment monitoring program and forensic sediment analysis at selected sites within Lake Macquarie, NSW; Final report on Monitoring Data. Connel Wagner PPI, Sydney, NSW Hunter Water (2011) Environmental Safeguards. Accessed 13 April. https://www.hunterwater.com.au/Water-and-Sewer/Wastewater- Systems/Environmental-Safeguards.aspx Jennings S, Warr KJ (2003) Environmental correlates of large-scale spatial variation in the δ15N of marine animals. Marine Biology 142:1131-1140 Keuskamp D (2003) Limited effects of grazer exclusion on the epiphytes of Posidonia sinuosa in South Australia. Aquatic Botany 78:14 Kousbrouek D, Raoult V, McKenzie L, Geary P, Gaston T (2017) Isotope analysis (15N and 13C) of autotrophs reveal the extent of anthropogenic nutrients in an urban estuary. Submitted Kumar IJN, Sajish PR, Kumar RN, Basil G, Shailendra V (2011) Nutrient Dynamics in an Avicennia marina (Forsk.) Vierh., Mangrove Forest in Vamleshwar, Gujarat, India. Notulae Scientia Biologicae 3:51-56

Recycled water in Cockle Creek Page 53

Leonard A, Steven A (2001) Water Quality objectives for marine and estuarine waters - ecosystem protection. EPA Victoria, Southbank, Victoria Lequerica M, McInnes R (2016) Evaluation of upgrade effects of four serwage treatment plants in NSW; Analysing the consequences of qulaitative and quantitative upgrades in sewage treatment. Water e-Journal 1 Lim S-R, Suh S, Kim J-H, Park HS (2010) Urban water infrastructure optimization to reduce environmental impacts and costs. Journal of Environmental Management 91:630-637 LMC Library (2017a) Edgeworth. Accessed 13 April. https://history.lakemac.com.au/page- local-history.aspx?pid=1085&vid=20&tmpt=narrative&narid=34 LMC Library (2017b) Salty Creek Recreation Area. Accessed 13 April 2017. https://history.lakemac.com.au/page-local- history.aspx?pid=1085&vid=20&tmpt=narrative&narid=4326 LMC Library (2017c) Waratah Golf Club. Accessed 13 April. https://history.lakemac.com.au/page-local- history.aspx?pid=1087&vid=20&tmpt=narrative&narid=3772 LMC Library (2017d) West Wallsend. Accessed 13 April. https://history.lakemac.com.au/page-local- history.aspx?pid=1085&vid=20&tmpt=narrative&narid=89 LMCC (2017) Fact Sheet 1: Environmental Facts & Impacts in Lake Macquarie. Living Lake Macquarie. Lake Macquarie City Council, Speers Point, NSW Logan JM, Jardine TD, Miller TJ, Bunn SE, Cunjak RA, Lutcavage ME (2008) Lipid corrections in carbon and nitrogen stable isotope analyses: comparison of chemical extraction and modelling methods. Journal of Animal Ecolocgy 77:838-846 Macko SA, Fogel ML, Hare PE, Hoering TC (1987) Isotopic fractionation of nitrogen and carbon in the synthesis of amino acids by microorganisms. Chemical Geology: Isotope Geoscience section 65:79-92 Manidis Roberts (2010) Incitec Fertilizers Ltd Cockle Creek Stage 2 Environmental Assessment. 2 - Appendices A to F Mariotti A, Lancelot C, Billen G (1984) Natural isotopic composition of nitrogen as a tracer of origin for suspended organic matter in the Scheldt estuary. Geochimica et Cosmochimica Acta 48:549-555 McKenzie L (2017) Email to T. Gaston. Melville AJ, Connolly RM (2003) Spatial Analysis of Stable Isotope Data to Determine Primary Sources of Nutrition for Fish. Springer-Verlag Merriam JL, McDowell WH, Tank JL, Wollheim WM, Crenshaw CL, Johnson SL (2002) Characterizing nitrogen dynamics, retention and transport in a tropical rainforest stream using an in situ 15N addition. Freshwater Biology 47:143-160 MHL (2002) Feasibility Study: Environmental improvements / Management of Cockle Bay. Manly Hydraulics Laboratory, Manly, NSW Mohan JA, Smith SD, Connelly TL, Attwood ET, McClelland JW, Herzka SZ, Walther BD (2016) Tissue-specific isotope turnover and discrimination factors are affected by diet

Recycled water in Cockle Creek Page 54

quality and lipid content in an omnivorous consumer. Journal of Experimental Marine Biology and Ecology 479:35-45 Moore K (2003) Uptake dynamics of oysters to detect sewage-derived nitrogen. Honours thesis, University of the Sunshine Coast, Sippy Downs, QLD Morrissey C, Boldt A, Mapstone A, Newton J, Ormerod S (2013) Stable isotopes as indicators of wastewater effects on the macroinvertebrates of urban rivers. Hydrobiologia 700:231-244 Murphy S (2006) Water quality and Water Cycle management report - Pasminco Cockle Creek Smelter. Maunsell Australia Py Ltd, Sydney, NSW Nixon SW (2009) Eutrophication and the macroscope. Hydrobiologia 629:5-19 Phillips DJ, Rainbow PS (1994) Definitions and Scope.In: Biomonitoring of trace aquatic contaminants Environmental Management Series, vol 37. Springer, Dordrecht Quiggin J (2001) Environmental economics and the Murray-Darling river system. Australian Journal of Agricultural & Resource Economics 45:67 Reis CRG, Nardoto GB, Rochelle ALC, Vieira SA, Oliveira RS (2017) Nitrogen dynamics in subtropical fringe and basin mangrove forests inferred from stable isotopes. Oecologia 183:841-848 Rodriguez C, Stoner AW (1990) The epiphyte community of mangrove roots in a tropical estuary: distribution and biomass. Aquatic Botany (Netherlands):117 Roy PS, Crawford EA (1984) Heavy metals in a contaminated Australian estuary—Dispersion and accumulation trend. Estuarine, Coastal and Shelf Science 19:341-358 Roy PS, Williams RJ, Jones AR, Yassini I, Gibbs PJ, Coates B, West RJ, Scanes PR, Hudson JP, Nichol S (2001) Structure and Function of South-east Australian Estuaries. Estuarine, Coastal and Shelf Science 53:351-384 Runcie JW, Ritchie RJ, Larkum AWD (2003) Uptake kinetics and assimilation of inorganic nitrogen by Catenella nipae and Ulva lactuca. Aquatic Botany 76:155-174 Rush J (2003) The nutirnet dynamics of an Australian river system as measured using stable isotopes. PhD thesis, University of Sydney, Sydney, NSW Ryall B (2005) Site Audit Report - Review of proposed groundwater remediation and monitoring Pasminco Cockle Creek Smelter site. HLA-Envirosciences Pty Ltd, Gordon, NSW Scanes P, Coade G, Doherty M, Hill R (2007) Evaluation of the utility of water quality based indicators of estuarine lagoon condition in NSW, Australia. Estuarine, Coastal and Shelf Science 74:306-319 Scanes PR, Philip N (1995) Environmental impact of deepwater discharge of sewage off Sydney, NSW, Australia. Marine Pollution Bulletin 31:343-346 Scanes PR, Scanes K, Otway NM (1995) Environmental problems due to disposal of wastes. In: Underwood AJ, Chapman MG (eds) Coastal marine ecology of temperate Australia. UNSW Press, Sydney, Australia

Recycled water in Cockle Creek Page 55

Sheaves M, Johnston R, Connolly RM (2012) Fish assemblages as indicators of estuary ecosystem health. Wetlands Ecology and Management 20:477-490 Signa G, Mazzola A, Kairo J, Vizzini S (2017) Small-scale variability in geomorphological settings influences mangrove-derived organic matter export in a tropical bay. Biogeosciences 14:617-629 Tan Y, Li J, Cheng J, Gu B, Hong J (2013) The sinks of dissolved inorganic nitrogen in surface water of wetland mesocosms. Ecological Engineering 52:125-129 Umwelt (2015) Lake Macquarie Coastal Zone Management Plan Part B For the Estuary. Lake Macquarie City Council, Teralba, NSW Vizzini S, Mazzola A (2003) Seasonal variations in the stable carbon and nitrogen isotope ratios (13C/12C and 15N/14N) of primary producers and consumers in a western Mediterranean coastal lagoon. Marine Biology 142:1009-1018 Watterson EK, Burston JM, Stevens H, Messiter D (2010) The hydraulic and morphological response of a large coastal lake to rising sea levels. Newcastle East, NSW Yamaguchi YT, Chikaraishi Y, Takano Y, Ogawa NO, Imachi H, Yokoyama Y, Ohkouchi N (2017) Fractionation of nitrogen isotopes during amino acid metabolism in heterotrophic and chemolithoautotrophic microbes across Eukarya, Bacteria, and Archaea: Effects of nitrogen sources and metabolic pathways. Organic Geochemistry 111:101-112 Zhen S, Zhu W (2016) Analysis of isotope tracing of domestic sewage sources in Taihu Lake— A case study of Meiliang Bay and Gonghu Bay. Ecological Indicators 66:113-120

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APPENDICES

APPENDIX A – COMPARISON OF PILOT STUDY & AUGUST RESULTS

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APPENDIX B – MISCELLANEOUS DATA

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