Artificially oxygenating the Swan River estuary increases dissolved oxygen concentrations in the water and at the sediment interface

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Article

Reference

Artificially oxygenating the Swan River estuary increases dissolved oxygen concentrations in the water and at the sediment interface

LARSEN, Sarah J., et al.

Abstract

The upper reaches of the Swan River estuary (Perth, Australia) has a history of eutrophication-related oxygen depletion, which has contributed to poor water quality and fish deaths. To alleviate hypoxic conditions, a trial side-stream supersaturation (SSS) oxygenation plant was established at Guildford (39 km upstream of the estuary mouth) in 2009. After notable success, a second plant was constructed at Caversham (44.2 km upstream of the estuary mouth) in 2011. Oxygenation plants have more commonly been used to treat deep, freshwater lakes and reservoirs and this is a pioneer application to a shallow estuary. We report on the effect of the Caversham plant on water and sediment condition over a 12-day experiment: before, during and post-plant operation. We monitored several physical and chemical parameters collected from daily longitudinal transects, moored continuous loggers, an acoustic Doppler current profiler and an in-situ sediment microprofiler. Oxygenation immediately improved dissolved oxygen concentrations in the water column and the distance over which the effect was observed was strongly influenced by the [...]

LARSEN, Sarah J., et al . Artificially oxygenating the Swan River estuary increases dissolved oxygen concentrations in the water and at the sediment interface. Ecological Engineering , 2019, vol. 128, p. 112-121

DOI : 10.1016/j.ecoleng.2018.12.032

Available at:

http://archive-ouverte.unige.ch/unige:123148

Disclaimer: layout of this document may differ from the published version.

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Contents lists available atScienceDirect

Ecological Engineering

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

Artificially oxygenating the Swan River estuary increases dissolved oxygen concentrations in the water and at the sediment interface

Sarah J. Larsen

^a,^⁎

, Kieryn L. Kilminster

^a,b

, Alessandra Mantovanelli

^a

, Zoë J. Goss

^a

, Georgina C. Evans

^a

, Lee D. Bryant

^c

, Daniel F. McGinnis

^d

aDepartment of Water and Environmental Regulation, Government of Western Australia, Locked Bag 10, Joondalup DC 6919, Western Australia, Australia

bSchool of Biological Sciences, University of Western Australia, 35 Stirling Hwy, Crawley 6009, Western Australia, Australia

cWater, Environment and Infrastructure Research (WEIR) Group, Department of Architecture and Civil Engineering, University of Bath, Claverton Down, Bath BA2 7AY, United Kingdom

dAquatic Physics Group, Department F.-A. Forel for Environmental and Aquatic Sciences (DEFSE), Faculty of Sciences, University of Geneva, Geneva, Switzerland

A R T I C L E I N F O Keywords:

Artificial oxygenation Hypoxia

Swan River estuary Side-stream supersaturation Sediment-water interface Oxygen flux

A B S T R A C T

The upper reaches of the Swan River estuary (Perth, Australia) has a history of eutrophication-related oxygen depletion, which has contributed to poor water quality and fish deaths. To alleviate hypoxic conditions, a trial side-stream supersaturation (SSS) oxygenation plant was established at Guildford (39 km upstream of the estuary mouth) in 2009. After notable success, a second plant was constructed at Caversham (44.2 km upstream of the estuary mouth) in 2011. Oxygenation plants have more commonly been used to treat deep, freshwater lakes and reservoirs and this is a pioneer application to a shallow estuary. We report on the effect of the Caversham plant on water and sediment condition over a 12-day experiment: before, during and post-plant operation. We monitored several physical and chemical parameters collected from daily longitudinal transects, moored continuous loggers, an acoustic Doppler current profiler and an in-situ sediment microprofiler. Oxygenation immediately improved dissolved oxygen concentrations in the water column and the distance over which the effect was observed was strongly influenced by the hydrodynamics of the estuary. After five days of oxygenation, water column dissolved oxygen had increased over a distance in excess of 11.5 km. In addition, oxygenation improved dissolved oxygen concentrations at the sediment-water interface, thereby increasing oxygen fluxes into the sediment. Ultimately, artificially oxygenating the Swan River estuary provides a refuge for fauna while facilitating aerobic decomposition of organic matter and enhancing nutrient cycling at the sediment-water interface. In light of the increasingly critical state of urbanised estuaries world-wide, results from this study highlights 1) the effectiveness of oxygenation in improving water quality and its potential for facilitating ecosystem restoration, and 2) the diversity of environments in which artificial oxygenation can be applied.

1. Introduction

Oxygen depletion in coastal and estuarine waters is a growing global concern, often symptomatic of excessive nutrient input from anthropogenic sources (Conley et al., 2009; Paerl, 2006). Hypoxia, defined here as dissolved oxygen (DO) concentrations less than 4 mg L⁻¹, occurs when oxygen is consumed (e.g. by decomposing organic matter, respiration and oxidation of reduced chemical species) faster than it is replenished (e.g. via air-water oxygen transfer, photosynthesis, and mixing) (Middelburg and Levin, 2009; Paerl, 2006). Highly- stratified estuaries, in which surface and bottom waters do not mix, are

more prone to hypoxia (Conley et al., 2009; Park et al., 2007). This condition is exacerbated when sediments are organic-rich and impose high sediment oxygen demands as is frequently the case in estuarine environments (Higashino et al., 2004; Valdemarsen et al., 2014).

Hypoxia has detrimental effects on aquatic fauna and vegetation and can significantly alter biogeochemical processes (Mascaro et al., 2009; Middelburg and Levin, 2009; Wu, 2002). Severe anoxic events can deplete the benthic invertebrate community (Tweedley et al., 2015), as well as result in large fish mortalities (Thronson and Quigg, 2008). Anoxia at the sediment-water interface promotes the release of reduced chemical species (e.g. nutrients and trace metals) which may

https://doi.org/10.1016/j.ecoleng.2018.12.032

Received 22 June 2018; Received in revised form 21 December 2018; Accepted 27 December 2018

⁎Corresponding author.

E-mail addresses:Sarah.Larsen@dwer.wa.gov.au(S.J. Larsen),Kieryn.Kilminster@dwer.wa.gov.au(K.L. Kilminster),

Alessandra.Mantovanelli@dwer.wa.gov.au(A. Mantovanelli),Zoe.Goss@dwer.wa.gov.au(Z.J. Goss),Georgina.Evans@dwer.wa.gov.au(G.C. Evans), L.Bryant@bath.ac.uk(L.D. Bryant),Daniel.Mcginnis@unige.ch(D.F. McGinnis).

Ecological Engineering 128 (2019) 112–121

Available online 15 January 2019

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be toxic to biota (e.g. ammonia and hydrogen sulphide) and confound ecological responses to low DO (Wu, 2002), or stimulate algal growth by increasing bioavailable nutrient concentrations (Bryant et al., 2010a;

Conley et al., 2009; Howarth et al., 2011).

Artificial oxygenation is an engineered means of supplementing oxygen demand. Most applications have been in lentic systems, in particular, deep freshwater lakes and reservoirs (Beutel and Horne, 1999; Gerling et al., 2014). In shallow water bodies oxygen dissolution within the water column is limited, making oxygenation techniques such as bubble-plume and line diffusers or submerged contact chambers less effective (Gerling et al., 2014; Singleton and Little, 2006). Side- stream supersaturation (SSS) plants overcome this challenge (Gerling et al., 2014; Sherman et al., 2012; Speece, 1996) by pumping water to a dissolution device and supersaturating with pure oxygen prior to dis- charging back into the waterbody for rapid dilution, thereby minimising bubble formation and losses to the atmosphere (Singleton and Little, 2006). While side-stream oxygenation was devised in the 1950s (Fast et al., 1975) and has applicability to shallow (< 10 m) water bodies there are only a few such applications in the literature. SSS has been successful in freshwater Falling Creek reservoir, Virginia (mean depth ∼4 m) (Gerling et al., 2014) and the Canning River weir pool, Western Australia (mean depth ∼2 m) (Greenop et al., 2001), however to our knowledge there are no published examples of the ongoing SSS oxygenation of an estuary (Gerling et al., 2014).

Here we describe a pioneer SSS oxygenation plant at Caversham, on the upper Swan River estuary, Western Australia. We investigate how estuarine hydrodynamics influence the distance added oxygen is car- ried up and downstream. We examine the impact of commencing and ceasing plant operation on water column DO as well as DO profiles at the sediment-water interface. We also share the implications of these finding to key design elements of SSS oxygenation plants when applied to shallow estuaries.

2. Methods 2.1. Study site

The Swan River estuary is a diurnal, micro-tidal estuary which abuts the city of Perth and discharges to the Indian Ocean at Fremantle in Western Australia (Fig. 1a). Circulation and mixing are driven by freshwater discharge and tidal forcing. The upper estuary is a narrow, shallow, and meandering channel (< 100 m wide, mean depth < 3 m):

generally well-mixed during winter and partially mixed to highly stratified in summer and autumn, depending on the inland migration of

the tidally-driven salt wedge and the relative intensity of tidal mixing and gravitational circulation (O'Callaghan et al., 2007). The maximum excursion of the salt wedge due to the diurnal tide was estimated to be 1.2 km (with a mean amplitude of 0.6 m) byO'Callaghan et al. (2007), propagating as a progressing wave from the mouth to the upper estuary in approximately 2.5 h. In addition, larger excursions of up to 5 km were measured as a result of sub-tidal water level changes which are remotely forced by continental shelf waves and storm surges and cause large shifts in circulation, salinity and DO.

The upper Swan River estuary has been characterised by high phytoplankton biomass and periodic hypoxia in the bottom waters (Hamilton et al., 2001). In 2009, the first ‘permanent’ oxygenation plant was constructed on the Swan River at Guildford, 39 km from the estuary mouth, to mitigate low oxygen conditions. The Guildford oxygenation plant was shown to improve DO concentrations in a 2010 investigation (Fig. S1, supplementary material), warranting the construction of a second plant, 5.3 km further upstream at Caversham in 2011 (Fig. 1b). This study focuses on results from a 12-day study of the Caversham plant performed in 2012.

2.2. Oxygenation plant

The Caversham oxygenation plant draws water at a rate of 120 L s⁻¹ from ∼0.4 m above the river bed via a self-priming centrifugal pump.

The intake structure was designed to minimise inflow velocities and prevent entrainment of debris and biota, as well as draw higher density water that would remain near the bottom when oxygenated. Liquid oxygen is stored in a pressurised vessel with a capacity of 14.2 m³at cryogenic temperatures. Liquid oxygen is vaporised to gas and com- bined with the extracted estuary water in a BOC Limited proprietary dissolver, operating at a pressure of 320 kPa, resulting in supersaturated oxygen concentrations of ∼110 mg L⁻¹. Treated water is returned to the river via a single pipe that splits into two 4.2-m sparge bars in a linear array, positioned ∼0.4 m above the river bed; each with 11 sharp-edged orifices on each side, orientated up and downstream. The orifices are sized to maintain adequate back pressure in the pipework and retain DO in solution, in addition to generating turbulent jets that are rapidly entrained by the receiving water body and diluted. Mixing of the supersaturated water in the estuary occurs via advection and diffusion. The oxygen flow rate can be varied from 10 to 60 kg h⁻¹in response to in-situ demands. Remote access to the plant control panel facilitates easy modification of the operating mode and the oxygen flow rate. Most commonly, the plant operates in ‘DO control’, responding to triggers from moored monitoring stations directly up and downstream

Fig. 1.(a) The Swan River estuary showing the location of the study site (star). (b) The locations of the two oxygenation plants (stars), transect sampling sites in the target impact zone (black dots) and sites further afield (grey dots). (c) The Caversham plant, its suction, and discharge lines, upstream (US) and downstream (DS) moored monitoring stations, the ADCP, and the sediment microprofiler.

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of the plant by switching on and off according to defined thresholds for low (e.g. 5 mg L⁻¹) and high (e.g. 9 mg L⁻¹) DO; however the plant can also operate in manual ‘on’ or ‘off’ modes or according to a timer.

2.3. Experimental design and data collection

The experiment commenced on 20 April 2012, and pre-operation data (3 d) were collected prior to turning the plant on at an oxygen flow rate of 40 kg h⁻¹ at 17:58 on 22 April. The plant ran continuously during the ‘on’ phase of the experiment (5 d) and was switched off at 15:28 on 27 April after which post-operation data (4 d) were collected until the conclusion of the experiment on 1 May 2012. The duration of

‘on’ and ‘off’ sampling periods was based on previous observations of response times to oxygenation and the subsequent drawdown of DO in the vicinity of the plant. The Guildford oxygenation plant was not in operation during the experiment.

One of the complexities of determining the impact of the oxygenation plant on the estuary is that the receiving water body is constantly moving and the direction and magnitude of flow are variable. A ‘target impact zone’ was designated from 3.1 km downstream to 3.2 km upstream (MEA to MSB inFig. 1b), which aligned with long-term monitoring sites within the design treatment area (based on estimates of tidal excursion; Department of Water and Environmental Regulation, unpublished data). Vertical water column profiles of pH, DO concentration and saturation, salinity and temperature were measured daily (between 7:48 and 14:19) with a Hydrolab DS5X multiprobe at sampling sites along a transect (Fig. 1b). Measurements were obtained at a depth of 0.2 m, then 0.5 m-increments from 0.5 m below the surface to 0.2 m above the river bed. Profiling was undertaken at approximately the same time of day and on a flooding tide to reduce the sampling variability associated with diurnal fluctuations and the direction of flow. The transect initially spanned the target impact zone (20 – 24 April), but it became apparent during the experiment that the cumulative effects of ongoing oxygenation was influencing further afield and the transect was extended to ∼4.8 km downstream to ∼6.7 km upstream of the plant from 25 April onwards (SUC to JET inFig. 1b).

Two continuous monitoring stations were moored 120 m upstream and 160 m downstream of the plant discharge (Fig. 1c) to 1) measure the impact of oxygenation irrespective of the orientation of the tide and 2) capture diurnal fluctuations not evident in daily transects. Mounted horizontally on each station were two Hydrolab DS5X multiprobes, one positioned 0.5 m below the surface and the other 0.5 m above the river bed, logging DO concentration and saturation, salinity, temperature, pH, and depth every ten minutes.

Current velocities were measured every five minutes by an upward- looking acoustic Doppler current profiler (ADCP; Teledyne Express Sentinel 1200 kHz) moored 145 m downstream of the plant discharge (within the vicinity of the sediment profiler, Fig. 1c). The ADCP blanking distance and bin sizes were ∼0.6 m and 0.25 m respectively, averaging 250 pings per ensemble. ADCP data were corrected for magnetic declination and quality controlled by removing data con- taminated by side-lobe interference or with low signal-to-noise ratio.

Edited data were placed into uniform depth strata, taking into account variations in the sampling depths with tides.

Sediment oxygen microprofiles were collected using an in-situ au- tonomous profiler (MiniProfiler MP4, Unisense A/S) fitted with two oxygen microsensors (OX-100, Unisense A/S) with a sampling fre- quency of 1 Hz. The sediment profiler was deployed 175 m downstream of the plant (Fig. 1c), at a site outside of the navigation channel with sufficient water depth (3.3–4.0 m) and easy access to a secure location onshore for housing the programming and storage component of the system (ARCTICA). The microsensors were Clark-type with a tip dia- meter of 100 µm, a fast response time (90% in less 8 s) and negligible stirring sensitivity. As the highly turbid nature of the upper Swan River estuary prevents visual inspection of the sensor tip height, the location of the sediment-water interface was initially estimated from profile

data. The linear section of the profile is associated with the diffusive boundary layer, and the depth at which there is a marked change in slope can be attributed to the porosity difference between water and sediment and hence the location of the sediment-water interface (Bryant et al., 2010a).

DO profiles (90 mm above to 40 mm below the sediment) were obtained as follows: 3 measurements at 90 mm above the sediment, followed by measurements at a resolution of 0.2–2 mm (3 per depth) from 50 mm above to 40 mm below the sediment-water interface, with high-precision measurements (5 per depth) collected at a resolution of 0.1 mm in the immediate range (5 mm above to 6 mm below) of the sediment-water interface. Each profile was completed in 78 min and referenced by the date and time of the first reading. Logged data were downloaded and batteries exchanged every 24 h, at which time mod- ifications were made to the program to realign the high-resolution portion of the profile with the location of the sediment-water interface, accounting for the distance the profiler had sunk into the sediment (up to 45 mm over the full 12-day deployment). A two-point calibration of the sensors was undertaken immediately before and after deployment, using a zero-oxygen solution (0% DO) of sodium ascorbate and NaOH (concentration of 0.1 M) and an atmospheric solution that has been bubbled vigorously with air (100% DO) in a calibration chamber.

2.4. Data analysis 2.4.1. Tidal migration

Tidal migration is defined here as the horizontal distance a water parcel moves during a diurnal tidal cycle, calculated for each depth strata by integrating the along-channel velocities over time (as measured by the ADCP). When summed over consecutive tidal cycles, the cumulative tidal migration estimates the maximum extent of influence of the Caversham oxygenation plant over the analysed period.

Unfortunately, current data are not available for the 24 April due to public interference requiring repositioning of the ADCP. Therefore, cumulative tidal migration was estimated for the periods between 20 − 23 April and 25 April − 1 May.

2.4.2. Vertical mixing

The strength of vertical mixing through the water column was estimated using a simplified layer Richardson number. Vertical mixing depends on the relative strength of factors generating turbulence (e.g.

velocity shear) and those effecting stratification (e.g. buoyancy). The Richardson number assumes mixing is dominated by velocity shear produced at the bottom (Dyer and New, 1986):

= Ri gh

L U2 (1)

whereRi_Lis the Richardson number,gis the acceleration due to gravity (m s⁻²),his the water depth (m), Δρis the difference between the bottom and surface densities (kg m⁻³), U is the depth-averaged current magnitude (m s⁻¹) and the depth-averaged density (kg m⁻³). For RiL> 20, bottom-generated turbulence is ineffective at decreasing stratification (vertical mixing inhibited, stable stratification); for 20 >RiL> 2 mixing is increasingly active (partially mixed, instable stratification) and for RiL≤ 2 mixing is fully developed (vertically mixed) (Cavalcante et al., 2013; Dyer and New, 1986).

2.4.3. Oxygen content

Total oxygen content and volume-weighted oxygen concentration were calculated for the target impact zone based on the method of Gantzer et al. (2009). Bathymetric data were used to divide the volume of the target impact zone into cells, each associated with a discrete DO concentration measured at its mid-point of depth and position along the transect. Total oxygen content was determined by summing the pro- ducts of each DO concentration measurement and its representative

S.J. Larsen et al. Ecological Engineering 128 (2019) 112–121

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volume across the target impact zone:

= ×

= =

Mass DO Vol

x s

z n

x z x z

1 1

, ,

(2) where Mass= total volume-weighted oxygen mass (kg), DOx,z= DO concentration at site (x) and depth layer (z),Volx,z= cell volume corresponding to site (x) and depth layer (z), s= number of sites, and n= number of layers. As the water depth varied across days and over the duration of sampling along the transect, a water surface elevation of 0.35 m (relative to Australian Height Datum) was used in calculations.

While total oxygen content is a simplified indicator of complex oxygen dynamics and does not quantify the individual components of supply and demand (e.g. wind mixing, photosynthesis or respiration), it is useful for comparison to the mass of oxygen added by the plant. It is a conservative estimate of oxygen accumulation as a result of plant operation as some of the added oxygen may have been transferred beyond the target impact zone in this lotic system.

The total oxygen content for each site in the target impact zone was divided by its representative volume to determine the daily volume- weighted oxygen concentration (as perGerling et al., 2014), with the rate of change in volume-weighted oxygen concentration over the five days of continuous oxygenation calculated via regression analysis.

2.4.4. Diffusive oxygen flux

Diffusive oxygen flux was calculated from Fick’s First Law of Diffusion applied to the sediment side of the sediment profile (Bryant et al., 2010b; Rasmussen and Jørgensen, 1992):

=

J Ds C

O2 z (3)

whereby ^C_z was determined from regression analysis of the linear section of the oxygen profile immediately below the sediment-water interface (Dalsgaard et al., 2000). The diffusion coefficient of oxygen in sediment (Ds) accounts for the effect of tortuosity by multiplying the diffusivity of oxygen in pure water (Dw) by the volumetric sediment porosity. Gravimetric porosity and total organic content (obtained via laboratory analysis of the top 2 cm of 9 replicate sediment cores collected within 20 m of the sediment profiler) were used to derive volumetric porosity (φ) (Avnimelech et al., 2001). The average porosity (value of φ = 0.7) was representative of estuarine systems (Glud, 2008), although lower than that measured ∼5.4 km downstream (φ = 0.92, Department of Water and Environmental Regulation, unpublished data) reflecting the heterogeneity of sediments in the Swan River estuary (Smith et al., 2010).

The position of the sediment-water interface in each profile (defined as the zero height datum) was refined by identifying the depth at which the standard deviation of repeated measurements was minimised, reflecting the reduced turbulence that occurs when approaching the sediment-water interface (Bryant et al., 2010a). The tidal flux frequently led to an instable water column oxygen profile; hence, estimates of diffusive flux based on the waterside of the profiles were not applicable (Bryant et al., 2010b).

3. Results

3.1. Estuarine circulation and tidal migration

When oxygenating an estuary, it is critical to understand the magnitude and direction of the current velocity as these factors determine if the supersaturated plume will be advected up or downstream, over what distance it will be advected, and the degree to which it will be diluted by the entraining river flow. For example, flooding tides will entrain DO in the discharge plume and advect it upstream, exposing upstream sites to recently oxygenated waters; while downstream sites would not have been exposed to oxygenated waters since the previous

ebb (∼12 h earlier). The converse will be true on the ebbing tide.

The along-channel velocity varied with depth (measured 145 m downstream of the plant discharge, seesupplementary material Fig.

S2), and consequently, the cumulative tidal migration varied in both magnitude and direction among different depth strata (Fig. 2a). Be- tween 20 – 23 April water movement was downstream in all water column depth strata, except near the bottom (∼0.6 m above the river bed) where the net movement was almost null (Fig. 2a, left panel). The cumulative tidal migration increased with proximity to the water surface. From 25 April – 1 May, water moved downstream in the surface and mid layers (∼1.6–2.6 m above the river bed) with a maximum migration of 12.9 km near the surface (Fig. 2a, right panel). Conversely, water movement was upstream near the bottom (∼0.6–1.4 m above the river bed) with a maximum migration of over 7 km.

Bottom waters were forced upstream by abnormal water levels, i.e.

∼0.46 m higher than the predicted tides (Fig. 2b, Barrack Street, Perth, Department of Transport, www.transport.wa.gov.au). The sub-tidal oscillations (obtained by applying a low-pass filter to water-level data to remove signals with a period of less than 36 h) were the result of meteorological forcing as indicated by a drop in barometric pressure (average of 1009 hPa on 29 April; Guildford climate station, Depart- ment of Water and Environmental Regulation,http://wir.water.wa.gov.

au) and localised rain (29.6 mm, Whiteman Park rain gauge, Bureau of Meteorology,www.bom.gov.au). The upstream migration of the salt wedge resulted in increased bottom salinity near the oxygenation plant and density stratification from 27 April onwards, as identified by an increase in theRiL(although daily fluctuations were large) (Fig. 2c).

3.2. Temporal and spatial patterns in oxygen

Water column DO concentrations measured directly upstream of the plant (190 m upstream, CAV) were strongly influenced by plant operation (Fig. 3). Prior to oxygenation, low DO concentrations (1.2–2.3 mg L⁻¹) were observed throughout the vertical profile (Fig. 3a). Oxygenation resulted in immediate enhancement of DO concentrations, with incremental increases with continued operation, such that by 27 April, DO concentrations through the water column ranged from 6.8 to 9.3 mg L⁻¹. In turn, bottom DO concentrations rapidly declined when the plant was switched off and four days post- operation, concentrations had decreased to less than 0.7 mg L⁻¹near the bottom while surface DO concentration remained elevated at 8.9 mg L⁻¹. The upstream monitoring station (120 m upstream of the plant discharge) measured an increase in bottom water DO on each flooding tide in excess of 4 mg L⁻¹, when the plant was operating (Fig. 3b). One tidal cycle after ceasing operation, a linear rate of decline in bottom water DO concentration of 0.93 mg L⁻¹d⁻¹(R²= 0.94) was observed.

The separation in surface and bottom DO concentrations at this time can be explained by the presence of stable stratification, in agreement with higher Richardson numbers (Fig. 2c).

Vertical profiles of DO concentration and salinity are shown along the estuary channel before, during and post-plant operation inFig. 4.

The bold, vertical grey line indicates the location of the plant discharge.

While the plant was operating, DO concentrations increased incrementally at all sites in the target impact zone (∼3.1 km downstream to ∼3.2 km upstream) with the largest increases observed upstream. In addition, the impact was apparent further afield; with well-oxygenated conditions from ∼4.8 km downstream to 6.7 km upstream after five days of continuous operation (Fig. 4c). A rapid decline in DO concentrations near the bottom is evident at all sites once oxygenation ceased.

Estuarine conditions were well-mixed prior to plant operation (Fig. 4f), partially-mixed during operation (Fig. 4h), and strongly stratified four days after operation ceased (Fig. 4j). Stratification pre- vented vertical mixing, evident in the alignment of DO (Fig. 4e) and salinity contours (Fig. 4j), and the contrast between high DO surface waters and anoxic bottom waters.

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The total oxygen content within the target impact zone increased linearly with plant operation at a rate of 701 kg d⁻¹(R²= 0.96) relative to the plant’s oxygen flow rate of 960 kg day⁻¹(Fig. 5), which equates to a volume-weighted increase of 1.0 mg L⁻¹d⁻¹. In the four days post-operation, the oxygen content declined at a rate of 301 kg d⁻¹(however the linear regression had a much lower R²value of 0.57) (Fig. 5), equating to a volume-weighted decline of 0.44 mg L⁻¹d⁻¹ across the target impact zone. A simple mass balance would suggest that daily oxygen addition by the plant (960 kg d⁻¹) minus the rate of oxygen accumulation (701 kg d⁻¹) would equal the rate of oxygen consumption (calculated to be 259 kg d⁻¹). Note that in this context,

‘accumulation’ and ‘consumption’ are both ‘net’ terms, i.e. the result of all of the oxygen generating and consuming processes occurring in the system (e.g. wind mixing, photosynthesis, chemical oxidation, and biological respiration). Assuming these processes were consistent across the experiment, calculated rates of oxygen consumption would be equal to the rate of depletion. The similarity between these numbers (259 and 301 kg d⁻¹) gives us confidence that calculated rates are indicative of the major consumptive processes of the system at the time of sampling.

Volume-weighted DO concentrations increased over the five days of plant operation at all sites in the target impact zone, however, the rate of change varied by site (Fig. 6). Sites with larger rates of change in volume-weighted DO are more strongly influenced by plant operation on a flooding tide. Upstream (positive distance inFig. 6), the rate of change in DO concentration increased from 1.1 mg L⁻¹d⁻¹at 0.19 km (CAV) to a maximum of 1.7 mg L⁻¹d⁻¹at 1.7 km (REG), beyond which the rate declined to 1.1 mg L⁻¹ d⁻¹ at 3.2 km (MSB). Downstream

(negative distance inFig. 6), the rate of change in volume-weighted DO concentrations declined from 1.2 mg L⁻¹ d⁻¹ at 0.16 km (WBRP) to 0.54 mg L⁻¹d⁻¹at 3.1 km (MEA). We extrapolated the linear portions of the plot to thex-axis, and estimated the influence of the plant to extend 6.2 km upstream and 5.9 km downstream of the plant discharge, i.e. well beyond the target impact zone.

3.3. Sediment oxygen profiles and diffusive oxygen flux

DO profiles were obtained nearly continuously at the sediment- profiler site located 175 m downstream of the plant discharge. InFig. 7, DO profile data characterising 6 mm above to 6 mm below the sediment–water interface, are compared two days prior to plant operation (20 – 21 April), after three days of operation (25 – 26 April) and three days post-operation (30 April). In general, DO concentrations increased across the sediment–water profile when the plant was operating with a larger oxic zone present within the sediment. Variability is high in profiles collected during the period of plant operation.

The average DO concentrations measured at the sediment–water interface increased when the plant was operating and declined when operation ceased (0.4 mg L⁻¹two days prior to operation, 2.2 mg L⁻¹ after three days of operation and 0.2 mg L⁻¹three days post-operation;

Fig. 8a). Average sediment oxygen flux followed the same pattern (0.6 mmol m⁻²d⁻¹two days prior to operation, 11.8 mmol m⁻²d⁻¹ after three days of operation and 0.8 mmol m⁻²d⁻¹three days post- operation;Fig. 8b).

Fig. 2.(a) Cumulative migration distance upstream (US) and downstream (DS) calculated for each depth strata, labelled by the distance from the river bed to the centre of the depth strata, data from 20 – 23 April shown in left panel and data from 25 April – 01 May shown in right panel. (b) Measured and predicted water levels at Barrack Street, Perth, and low-pass filtered water levels (to remove signals with a period of less than 36 h). Water levels are relative to low water mark at Fremantle (0.76 m below Australian Height Datum) (c) The layer Richardson number (RiL); horizontal dashed lines delineate regions where:RiL> 20, bottom-generated turbulence appears ineffective at decreasing stratification (vertical mixing inhibited), for 20 >RiL> 2, mixing is increasing active (partially mixed), and for RiL≤ 2, mixing is fully developed (vertically mixed). The vertical dotted lines indicate when plant operation commenced and ceased on 22 and 27 April, respectively.

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4. Discussion

4.1. Hydrodynamics affects oxygen dispersion

The stretch of the estuary that will be influenced by the oxygenation plant is highly variable and dependent on a number of factors including tidal forcing, river flow, mixing and internal rates of oxygen consumption by biota and organic matter decay. The cumulative migration of bottom waters (either up or downstream) during the period of plant operation can be considered as the maximum reach of plant influence.

The effect of the plant will be evident in a subset of this range, as entrained DO will be consumed as it moves up and downstream. The faster the current, the larger the distance in which DO may be advected, and the greater the dilution of the oxygenated plume. In the latter part of the experimental period (25 April – 1 May), sub-tidal oscillations led to a larger intrusion of the salt wedge, with a cumulative upstream migration of the bottom waters (∼1 m from the river bed) in excess of 7 km. Over the same period, surface waters were more strongly influenced by river flow, particularly post-rainfall event, with a cumulative migration downstream of 12.9 km.

The influence of the plant was greater upstream over the sampled period, possibly due to: 1) water column profiling only on a flooding tide, 2) DO entrainment in bottom waters which had a cumulative

migration upstream, and 3) higher sediment oxygen demands downstream (Smith et al., 2007) may consume added DO more rapidly – also supported in our own data with the more rapid drawdown of DO in the bottom waters of downstream sites post-operation. Our empirical results have since been supported by three-dimensional modelling of the Swan estuary, which showed oxygenated water to have net migration upstream (Huang et al., 2018).

Downstream of the plant, the rate of increase in volume-weighted DO concentration was higher the closer the site was to the point of plant discharge, however upstream, the peak rate of increase was observed at a distance of 1.7 km, with a decline further afield. There are several potential explanations for why the peak oxygen increase upstream was not observed in the immediate location of the plant. It may be that it takes some distance for the oxygenated plume to mix laterally through the cross-section of the estuary, and water column profiling sites in close proximity to the plant may be positioned on the peripheral of the mixing zone where concentrations are not as high. Additionally, during the experimental period, there was a net migration of the tide upstream, potentially concentrating oxygen in this region. Alternatively, another oxygen source such as photosynthesising phytoplankton may be con- tributing to the peak in DO concentrations 1.7 km upstream but it is impossible to disentangle these factors in our study.

While the focus of this experiment was the target impact zone Fig. 3.(a) Dissolved oxygen (DO) concentrations (mg L⁻¹) measured at discrete depths through the water column (crosses) 190 m upstream of the plant discharge (CAV), with contours interpolated (kriging) between sampling days. (b) Continuous probe measurements of DO concentration (mg L⁻¹) measured 0.5 m below the surface and 0.5 m above the river bed (120 m upstream of the plant discharge). An increase or decrease in water depth (m) indicates a flooding and ebbing tide, respectively. The linear rate of decline in bottom water DO concentration was 0.93 mg L⁻¹d⁻¹, calculated from the next flooding tide after operation ceased (dashed line, R²= 0.94). In both (a) and (b) the period beetween vertical lines indicates when the plant was operating.

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Fig. 4.Dissolved oxygen (DO; mg L⁻¹) and salinity profiles along the estuary channel (∼4.8 km downstream and ∼6.7 km upstream the plant site): (a, f) one day prior to plant operation, (b, g) after one day of plant operation, (c, h) after five days of plant operation, (d, i) one day after operation ceased, and (e, j) four days after operation ceased. The bold, vertical grey line indicates the location of the plant discharge, with sites CAV – JET upstream and sites WBRP – SUC downstream.

Fig. 5.Total oxygen content in the target impact zone (∼3.1 km downstream to

∼3.2 km upstream) pre and post-operation (hollow dots) and during operation (solid dots) in comparison to the cumulative addition of oxygen by the plant (continuous grey line). Oxygen accumulated at a rate of 701 kg d⁻¹(R²= 0.96) and depleted at a rate of 301 kg d⁻¹(R²= 0.57). Vertical dotted lines indicate when the plant started and stopped respectively.

Fig. 6.The rate of change in volume-weighted dissolved oxygen (DO; mg L⁻¹ day⁻¹) after five days of continuous oxygenation calculated at profiling sites within the target impact zone ∼3.1 km downstream (negative distance) to

∼3.2 km upstream (positive distance). The vertical line indicates the location of the plant. The linear portions of the plot are extrapolated to thex-axis (dashed lines) to estimate the extent of plant influence on DO concentrations (6.2 km upstream and 5.9 km downstream of the plant discharge).

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(∼3.1 km downstream to ∼3.2 km upstream), we observed that sites further away were also influenced by oxygenation. The rapid decline in bottom water DO concentrations when operation ceased at sites as far as ∼4.8 km downstream and ∼6.7 km upstream is a clear demonstra- tion of the wide-reaching influence of the plant. A similar reach of influence was observed when we extrapolated the rate of change in volume-weighted DO concentration beyond the target impact zone and estimated that the influence of the plant extended from ∼5.9 km downstream to ∼6.2 km upstream after five days of operation. As theorised, the reach of plant influence is a subset of the cumulative migration distances calculated for surface and bottom waters of ∼7 km downstream and ∼12.9 km upstream.

The tidal advection of added DO upstream and downstream is a tangible benefit of oxygenating a tidal estuary in comparison to a reservoir or weir pool which would require long pipework to achieve the same distribution. While DO will also be advected in lentic systems, rates are likely to be much lower. For example, average seiche-induced currents of 1 cm s⁻¹have been shown to disperse added DO through the hypolimnion of Lake Hallwil, Switzerland (220 m in a 6 h period;

McGinnis et al., 2004), compared to average currents of 7.5 cm s⁻¹ measured in our study.Lorke et al. (2003) and Bryant et al. (2010a) measured maximum seiche-induced currents of 3 cm s⁻¹in the bottom waters of unoxygenated Lake Alpnach, Switzerland and similar currents were measured in oxygenated Carvins Cove Reservoir, Virginia (Bryant et al., 2011) in comparison to maximum currents of 26 cm s⁻¹in our study. Faster rates of advection will also more rapidly dilute the discharge jets and minimise localised oxygen supersaturation.

As the tidal excursion is variable, the oxygenation plant must have flexible oxygen flow rates and the ability to respond to real-time changes in estuary DO concentrations for optimal effectiveness. The Caversham oxygenation plant switches on automatically when either the upstream or downstream moored monitoring stations detects DO concentrations in bottom waters below a minimum threshold (adjustable, but generally set to 5 mg L⁻¹). This logic enables the plant to respond to incoming low-DO water irrespective of the direction of estuary currents. The plant switches off when both monitoring stations detect DO concentrations in bottom waters that are above the maximum threshold (adjustable, but generally set to 9 mg L⁻¹). The upper threshold is a balance between minimising supersaturation at the site of

discharge and maximising the amount of oxygen available for uptake when advected up and downstream.

Density stratification in the latter part of the experimental period influenced the advection and dispersion of the added oxygen. The alignment of DO and salinity contours (Fig. 4) indicates that the pyc- nocline acted as a physical barrier to the mass transport of oxygen across density layers (Kurup et al., 1998), a phenomenon that is often observed when artificially oxygenating. As such, residual DO persists in surface waters after operation ceased while the sediment oxygen demand rapidly depleted DO in bottom waters.

The Caversham oxygenation plant generally operates when the salt wedge is in the target impact zone, due to limited vertical mixing which is a controlling factor in DO drawdown (Huang et al., 2018). As such, the impact of stratification is an important consideration in plant design. The position and characteristics of the intake and discharge structures will determine the density of the treated water relative to that of the receiving water body. Drawing in water from close to the river bed will maximise its salinity (and density) and ensure that added DO remains near the sediment where it is available to supplement the sediment oxygen demand. It will also ensure that entrained DO can be advected up and downstream via the tidal excursion. When designing an oxygenation plant for a salt wedge estuary, the maximum reach of influence should be estimated from the tidal migration over single and multiple tidal cycles.

Fig. 7.In-situ dissolved oxygen (DO) profiles 6 mm above and 6 mm below the sediment-water interface two days prior to plant operation (20 – 21 April), after operating for three days (25 – 26 April) and three days post-operation (30 April).

Fig. 8.(a) Dissolved oxygen (DO) concentrations at the sediment-water interface (mg L⁻¹) (b) Diffusive DO flux into the sediment (mmol m⁻²d⁻¹). For both (a) and (b) the box represents the 25th and 75th percentiles, whiskers indicate the outlining data points, mean and median values are shown with a dotted and solid line respectively and the hashed block highlights the period when the plant was operating. The number of sediment profiles used to determine both box plots is indicated by the n value aligned with each date.

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4.2. Oxygenation increased water column DO

Volume-weighted DO concentrations increased at a rate of 1 mg L⁻¹ d⁻¹(Fig. 6) across the target impact zone when oxygenating and the system shifted from mostly hypoxic to well-oxygenated over the five days of operation. The residual DO was consumed quickly – the volume- weighted DO concentration declined at a rate of 0.44 mg L⁻¹day⁻¹. Four days post-operation the average DO concentration in the target impact zone had returned to essentially pre-operation levels.

DO concentrations increased throughout the water column but declined most rapidly closer to the sediment and at sites downstream of the plant discharge. Declining DO concentration with depth suggests that the sediment oxygen demand is a large component of the overall oxygen demand of the system, which is common in organic-rich and shallow systems such as estuaries, where there is less time for organic matter to biodegrade in the water column before settling to the sediment (Beutel et al., 2007; Higashino et al., 2004). Identification of high sediment oxygen demands emphasises the importance of directing supplementary oxygen to the bottom layers of the water column.

4.3. Oxygenation shifted oxygen dynamics at the sediment-water interface Artificial oxygenation resulted in a stepwise change in sediment oxygen profiles and an increase in oxygen penetration into the sediment. This enhanced sediment oxic zone reverted back to initial, anoxic conditions once operation ceased (Fig. 7). Plant operation increased DO concentrations at the sediment-water interface: average DO was 0.4 mg L⁻¹two day prior to operation, increased to 2.2 mg L⁻¹after three days of operation and decreased back to 0.2 mg L⁻¹three days post-operation. Diffusive oxygen flux into the sediment increased by an order of magnitude when comparing fluxes after three days of operation to those from two days prior to operation. The increase in the differential between DO concentrations in the sediment and the bottom waters when the plant is operating is a direct driver of increased diffusive oxygen fluxes into the sediment (Bryant et al., 2010a; Lorke et al., 2003).

The average diffusive oxygen flux (average flux of 11.8 mmol m⁻² day⁻¹) while oxygenating was less than that measured using similar microprofiler methodology in 2010, 5.4 km downstream at the Guildford oxygenation plant (average flux of 33.1 mmol m⁻²day⁻¹; unpublished data, Department of Water and Environmental Regulation). Measurements of sediment oxygen flux from benthic chambers in a similar region (∼5 km downstream) were also much higher (55–87 mmol m⁻² d⁻¹) (Smith et al., 2010). However, fluxes obtained in our Caversham 2012 study are of similar magnitude and are directly comparable to sediment-water fluxes obtained in other microprofiler-based studies (Bryant et al., 2011; Bryant et al., 2010a; Lorke et al., 2003). The lower fluxes measured in this study relative to those reported previously for the Swan River estuary may be due to variation in organic matter, bioturbation and/or mixing artefacts (in the case of the benthic chambers) and/or permeable sediments (McGinnis et al., 2014). The observed range of flux values highlights the variability of sediment oxygen demand, which occurs not only in the Swan River estuary, but in most natural systems (Bryant et al., 2010a; Glud, 2008).

Further field studies to characterise the variability of sediment oxygen demand in the Swan River estuary would enhance hydrodynamic-biogeochemical modelling (Huang et al., 2018) and improve our under- standing of what drives oxygen demand, and ultimately, water quality of the system as a whole.

If the shift from an anoxic to an aerobic sediment-water interface can be maintained, the impact on key biogeochemical processes and habitation by benthic fauna may be transformational. Phosphorus may be more strongly bound to iron in the sediment, and ammonium fluxes into the water column may be lower under aerobic conditions (Beutel, 2006; Beutel and Horne, 1999; Middelburg and Levin, 2009). Oxygen is required for nitrification, which supplies nitrate and nitrite for the nitrogen-removing processes of denitrification and anammox (Beutel and

Horne, 1999; Hietanen and Lukkari, 2007). The loss of benthic bio- turbators under anoxic conditions has direct consequences for nitrogen cycling (Howarth et al., 2011). Burrowing and tube-building by macro- and micro-fauna leads to irrigation of the sediment and distribution of DO into much deeper layers than possible by molecular diffusion alone (Glud, 2008). This increases the surface area exposed to the juxtapo- sition of anoxic and oxic conditions that support nitrification and denitrification cycles (Diaz and Rosenberg, 1995). The release of nitrogen and phosphorus to the water column under anoxic conditions represents a positive feedback loop; increases in available nutrients drives primary productivity which creates further oxygen demand as DO is depleted during eventual decomposition (Diaz and Rosenberg, 1995).

Prior to the construction of the two oxygenation plants at Caversham and Guildford, the upper Swan River estuary frequently exhibited extended periods of widespread hypoxia and anoxia (Department of Water and Environmental Regulation, unpublished data,http://wir.wwater.wa.gov.a). Artificial oxygenation has proven to be extremely effective in reducing the severity of low DO conditions and succeeds in off-setting the oxygen deficit almost year-round.

Further studies that look at long-term patterns in DO, as well as changes to fauna such as key benthic species, are warranted to truly appreciate the success of this intervention technique.

4.4. Artificial oxygenation can be applied to estuaries

When the Caversham oxygenation plant was constructed in 2011, it was not anticipated that it would be as effective as it was, nor that it would have an impact over such a large distance. While the rates of change in DO and the reach of plant influence will be highly variable, the empirical findings of this study reinforce the value of artificial oxygenation in the Swan River estuary. Building upon this study, the development of a coupled hydrodynamic-biogeochemical model provides further evidence that artificial oxygenation is able to make large- scale changes to the estuary over a wide range of environmental conditions and operational scenarios (Huang et al., 2018).

Oxygen conditions have deteriorated rapidly in many urbanised estuaries around the world as a result of eutrophication (Diaz and Rosenberg, 1995). Reducing the primary causes of nutrient enrichment should be the ultimate aim in the restoration of estuaries, allowing for a balance between oxygen supply and consumption (Conley et al., 2009).

However, while catchment-scaled management strategies are considered and applied, oxygenation plants can supplement oxygen demand, offering a refuge for fauna while facilitating aerobic decomposition of organic matter and enhancing nutrient cycles at the sediment-water interface. If appropriately designed, artificial oxygenation can immediately improve water quality with less institutional, technical and implementational constraints than catchment-wide strategies (Beutel et al., 2007; Gantzer et al., 2009).

5. Conclusions

In this study, the Caversham oxygenation plant was found to immediately improve oxygen conditions in the upper Swan River estuary, near Perth, Australia, with the magnitude and spatial extent of impact increasing incrementally over five days of operation. The extent of the plant’s influence is controlled by the hydrodynamics of the estuary and during the study we observed an increase in water column DO over a reach in excess of 11.5 km as a result of oxygenation. Locally, the plant substantially increased oxygen concentrations at the sediment–water interface, thereby supporting an increase in oxygen flux to the sediment.

This study shows for the first time that side-stream supersaturation (SSS) can be successfully applied to shallow estuaries where in-situ dissolution is often prohibitive, thereby fundamentally changing the oxygen dynamics of the system. With eutrophication pressures on estuaries increasing around the world, artificial oxygenation is a tool

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worthy of consideration in the restoration of these precious ecosystems.

Acknowledgments

The construction, operation, and monitoring of the Swan River oxygenation plants is a partnership between the Western Australian Department of Water and Environmental Regulation and the Department of Biodiversity, Conservation and Attractions. This project was funded by the Government of Western Australia and the Burswood Trust. The authors would like to thank the staff in the Aquatic Science Branch, Department of Water and Environmental Regulation for their field assistance and in particular Malcolm Robb for his tireless con- tributions to the project since its inception. The manuscript was greatly improved by the comments of an anonymous reviewer.

Appendix A. Supplementary data

Supplementary data to this article can be found online athttps://

doi.org/10.1016/j.ecoleng.2018.12.032.

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