Effect of temperature and drying-rewetting of sediments on the partitioning between denitrification and DNRA in constructed urban stormwater wetlands

Ecological Engineering 140 (2019) 105586

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Effect of temperature and drying-rewetting of sediments on the partitioning between denitrification and DNRA in constructed urban stormwater wetlands

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Moklesur Rahman , Michael R. Grace, Keryn L. Roberts, Adam J. Kessler, Perran L.M. Cook Water Studies Centre, School of Chemistry, Monash University, Clayton, Australia

A R T I C LE I N FO

A B S T R A C T

Keywords: Denitrification DNRA Wetlands Drying-rewetting Temperature

Constructed wetlands are increasingly used within urban areas to reduce pollutant runoff, including nitrogen. These environments are exposed to frequent wetting and drying events, and also increasing temperatures due to the heat island effect and seasonal variation. In this study, we have investigated the role of drying and rewetting of sediments and temperature on the partitioning of two nitrate reduction pathways: denitrification (DNF) and dissimilatory nitrate reduction to ammonium (DNRA) in four constructed urban stormwater wetlands in Melbourne, Australia. Our results suggest that both DNF and DNRA decreased due to drying of sediments, but DNRA decreased to a greater extent. Rates of DNRA from the drying-rewetting treatment decreased 5–90% compared to the control. Moreover, a concomitant increase in NH4+ fluxes and a slight change in sediment oxygen demand (SOD) suggest that oxic conditions arising from sediment drying was responsible for the decreased rates of DNF and DNRA. Rates of both DNF and DNRA increased with increasing temperature in slurries. The DNF:DNRA ratio suggests that the relative increase in DNRA was more than DNF with increasing temperature. The mean activation energy of DNF ranged from 41 ± 1 to 64 ± 4 kJ mol−1 and the corresponding temperature coefficient (Q10) values ranged from 1.3 to 2.4. In comparison, the mean activation energy of DNRA was higher and ranged from 50 ± 8 to 107 ± 14 kJ mol−1 with the corresponding Q10 values ranging from 1.2 to 3.4. Overall, our results suggest that drying and re-wetting of sediments decreases nitrate reduction and increases nitrogen retention, whereas increased temperature enhances the recycling of bioavailable nitrogen in wet sediments in constructed urban stormwater wetlands.

1. Introduction Constructed urban wetlands are treatment systems that use natural processes involving wetland vegetation, soils, and their associated microbial assemblages to improve water quality and are also considered effective to support wildlife (Reddy and DeLaune, 2008; Kadlec and Wallace, 2009). Constructed urban wetlands are typically far less expensive to build compared to traditional wastewater treatment plants and are a technically feasible approach to treating urban stormwater (Carlisle and Mulamoottil, 1991; Li et al., 2017). In urban areas, excess nutrients such as nitrogen (N) and phosphorus (P) from stormwater runoff, agricultural fertilizers and leaking septic tanks are removed in constructed wetlands. Wetlands are an interface between terrestrial and aquatic environments and hence are substantially influenced by the water cycle.

Freshwater wetlands are among the world’s most valuable ecosystems (Costanza et al., 1997; Zedler and Kercher, 2005; Gitay et al., 2011), yet they are expected to be affected by seasonal variabilities (Mesquita et al., 2017; Nhamo et al., 2017; Olivie-Lauquetet al., 2001; Poiani and Johnson, 1991; Poiani and Johnson, 1993; Larson, 1995; Sorenson et al., 1998; Johnson et al., 2005; Jin, 2008). Temperature is an important environmental factor that influences microbial respiration by influencing the reduction-oxidation (redox) conditions in wetlands (Rayner, 1998; Chen et al., 2000; Kadlec and Reddy, 2001; Chuankuan et al., 2002). Higher water column temperature in summer results in the decreased concentration of dissolved oxygen in water (Harvey et al., 2011). As a result, the hypoxic (O2 < 100 µmol L−1) or anoxic conditions in summer could be prolonged due to high respiration rates in the water column (Ficke et al., 2007). However, there could be a recovery of a portion of the depleted O2 through photosynthesis carried out by

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Corresponding author. E-mail addresses: [email protected] (M. Rahman), [email protected] (M.R. Grace), [email protected] (K.L. Roberts), [email protected] (A.J. Kessler), [email protected] (P.L.M. Cook). https://doi.org/10.1016/j.ecoleng.2019.105586 Received 29 January 2019; Received in revised form 20 August 2019; Accepted 25 August 2019 Available online 19 September 2019 0925-8574/ © 2019 Elsevier B.V. All rights reserved.

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facultative aerobic denitrifying bacteria to dry conditions, will lead to higher DNF than DNRA.

phytoplankton during the daylight hours (Malone, 1977; Harding et al., 1987; Eilers and Peeters, 1988). Hypoxic and anoxic conditions in constructed urban wetlands are conducive for nitrate reduction processes such as DNF and DNRA. Moreover, increased water temperature may influence the species composition of sediment microbial communities (Harder and Veldkamp, 1971; Nedwell and Floodgate, 1971; Inniss and Mayfield, 1978, 1979; Tison et al., 1980; Brin et al., 2015). For example, in sediments from the Colne estuary at the Hythe, Colchester, UK, a mesophilic DNRA bacteria (predominantly Pseudomonas sp.) was more active in summer than in winter (Ogilvie et al., 1997). In another study, higher water temperature during summer resulted in the highest rate of DNRA whereas the highest rate of DNF was observed during winter in coastal sediments (Jørgensen and Sørensen, 1988). The efficiency of DNRA bacteria to reduce nitrate is higher at higher temperatures and this will allow them to outcompete denitrifiers at higher temperature (Ogilvie et al., 1997). Kelly-Gerreyn et al. (2001) investigated the temperature effect on the partitioning between DNF and DNRA in the lower River Great Ouse sediments, Norfolk, UK through modelling and found that DNF was favoured within a narrow temperature range (14–17 °C) whereas DNRA was the dominant process at either side of this temperature window. From this result, it was concluded that temperature is an important factor controlling the partitioning of nitrate reduction between DNF and DNRA. However, there is still a lack of rigorous study on the effect of temperature on the partitioning of NO3− reduction between DNF and DNRA particularly in constructed urban wetlands. Vulnerability of wetlands due to seasonal changes (Barros and Albernaz, 2014) may arise through alterations to evapotranspiration and precipitation patterns (Grover, 2015). Higher temperatures in summer can subsequently cause drying of sediments in wetlands (Johnson et al., 2010). In particular in a country like Australia, wetland sediments are subjected to drying during summer due to the quick evaporation of the overlying water. Sediment bacterial community composition and rates of sediment biogeochemical processes are affected by the frequency of drying-rewetting of sediments (Fierer and Schimel, 2002). Denitrification (DNF) is a dissimilatory microbial process that converts NO3− to N2 and hence removes excess N from aquatic ecosystems to the atmosphere (Burgin and Hamilton, 2007; Reddy and DeLaune, 2008; Kadlec and Wallace, 2009). In contrast, dissimilatory nitrate reduction to ammonium (DNRA) reduces NO3− to NH4+ and thus, causes recycling of N within the ecosystem (An and Gardner, 2002; Burgin and Hamilton, 2007; Reddy and DeLaune, 2008; Kadlec and Wallace, 2009). Drying and rewetting influences the reductionoxidation (redox) by switching between oxic and anoxic conditions, and thus may affect both DNF and DNRA that occur in the anaerobic zone of sediments. The oxic conditions of sediments caused by drying would result in decreasing rates of both DNF and DNRA because of the oxygen inhibition of the enzymes involved in nitrate reduction (Wrage et al., 2001). Only a few studies have examined the effect of temperature and drying-rewetting of sediments (Venterink et al., 2002; Kjellin et al., 2007; Fromin et al., 2010) on DNF and DNRA (King and Nedwell, 1984; Ogilvie et al., 1997; Kelly-Gerreyn et al., 2001; Veraart et al., 2011). Therefore, very little is known about the effect of temperature and drying-rewetting of sediments on the partitioning of NO3− reduction between DNF and DNRA in constructed urban stormwater wetlands. This study investigates the response of N removal efficiency to temperature and drying-rewetting cycles. To investigate the effect of temperature on the partitioning of NO3− reduction between DNF and DNRA, we used four different temperatures in this study that reflect typical winter high and low and summer high and low temperatures. We hypothesised that nitrate ammonifying bacteria have higher affinity for substrates at higher temperature and thus, the increase in DNRA would be more than DNF with increasing temperature. We also tested the hypothesis that drying of the sediments will result in more oxic conditions in sediments, which together with the high resistance of

2. Materials and method 2.1. Sampling The four freshwater constructed urban wetlands in Melbourne, Australia used in this work are those previously used for studies on nitrate reduction processes (Huntingdale Road, Cascades on Clyde, Koolamara Blvd and Namatjira Reserve wetlands) (Rahman et al., 2019a,b). All these four wetlands are located in urban areas and commonly receive urban stormwater. The studied sites were dominated by Juncus sp. and to a lesser extent by Typha sp. However, the organic carbon (OC) content varied significantly among these wetlands (Rahman et al., 2019a,b). The Huntingdale Road wetland (37° 53′ 38″ S, 145° 6′ 39″ E) was constructed in early 2003 with an area of 18,500 m2 and is located on the floodplain of Scotchman’s Creek. The Cascades on Clyde wetland is located in North Clyde (38° 6′ 10″ S, 145° 19′ 37″ E) and was constructed in May 2008 with an area of 105,258 m2. This wetland services runoff from the adjacent urban and farming district. The Koolamara Blvd wetland (37° 53′ 50″ S, 145°16′ 32″ E) was constructed in 2005 and the area of the wetland is 22,461 m2. The Namatjira Reserve wetland is within parkland (Namatjira Park) at Clayton South (37° 56′ 5″ S, 145° 6′ 40″ E). The wetland was constructed in 2012 with an area of 45,000 m2 and services a mixed urban and industrial catchment. In situ water quality parameters (temperature, pH, DO, conductivity, turbidity) were measured at the inlet of each four wetlands using multi probe (Horiba U-50). Sediment cores for slurry experiments were collected by hand at the inlet of all four wetlands with polyethylene cylinders (6.5 cm inner diameter; 27–29 cm height). A summary of the sampling campaign and the experiments has been provided in Table 1. Site water was collected using 20 L carboys from the inlet of each wetland. Immediately on collection, cores were placed in the dark and on ice, and processed in the laboratory within 2 h. Sediment porewater was extruded from top 1 cm through centrifugation and filtered surface water samples for NH4+, FRP and NOx were collected by filtering through 0.2 µm PES filters (Sartorius). 2.2. Intact core incubations for drying experiments Rates of DNF and DNRA and nutrient fluxes were measured in intact sediment cores in the laboratory. DNF (D15) and DNRA (DNRA15) rates were measured from the reduction of added 15NO3− tracer. Overlying aerated site water column in cores were mixed using a magnetic stirrer (~40 rpm) suspended ~3–5 cm above the sediment surface to prevent disturbance of sediment and allowed to equilibrate overnight before experiments. The following day, fluxes of NH4+, NOx (defined as NOx = NO3− + NO2−), filterable reactive phosphorus (FRP) and dissolved oxygen (DO) were measured from three treatments of four replicate cores: wet, dry-rewet and control (Fig. 1) (Dalsgaard et al., Table 1 Summary of the sampling campaign for the temperature and drying-rewetting experiments.

2

Sampling period

Sampling sites

Type of experiment

Summer (Dec 2016) Summer (Dec 2016) Summer (Jan 2017) Summer (Jan 2017) Autumn (Mar 2017) Autumn (Apr 2017) Autumn (Apr 2017) Late autumn (May 2017)

Huntingdale Road Namatjira Reserve Koolamara Blvd Cascades on Clyde Huntingdale Road Namatjira Reserve Koolamara Blvd Cascades on Clyde

Temperature Temperature Temperature Temperature Drying-rewetting Drying-rewetting Drying-rewetting Drying-rewetting

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Fig. 1. Schematic of the intact core incubation method of measurement at time 1 (T1) repeated for other times T1–T4 for all treatments on day 1 and for control and rewet on day 9, a) measurement of DO in core 1, b) measurement of pH and DO in core 1 and 2, respectively, c) measurement of pH and DO in core 2 and 3, respectively and sample collection for nutrient fluxes from core 1 d) adding 15NO3− tracer in core 1 and series repeated for DO, pH and sample collection in core 2, 3, and 4 respectively. Cores were slurried from all treatments to measure DNF and DNRA at times T1-T4 after collecting samples for nutrient fluxes and adding tracer.

Table 2 Physico-chemical parameters measured at time of each sampling from all four wetlands from December 2016 to May 2017. Sampling location

Sampling dates

Temp. (0C)

pH

%DO saturation

Conductivity (mScm−1)

Turbidity (NTU)

Surface water

Porewater

NH4+ (µM)

NOx (µM)

FRP (µM)

NH4+ (µM)

NOx (µM)

FRP (µM)

Huntingdale Road

Dec 2016 Mar 2017

23.8 21.8

6.9 6.6

61.5 80.0

0.5 0.2

15.0 28.0

27.0 5.0

0.3 31.1

0.3 0.5

701 ND

0.0 ND

3.5 ND

Namatjira Reserve

Dec 2016 Apr 2017

22.3 17.4

8.9 6.7

70.0 81.0

0.3 0.2

8.6 15.0

0.9 2.2

23.2 58.1

0.5 0.6

442 260

0.0 0.4

0.6 0.7

Koolamara Blvd

Jan 2017 Apr 2017

27.7 15.4

7.8 6.8

113.0 103.0

1.1 0.7

5.8 16.0

0.7 1.5

1.0 114.1

0.4 0.4

251 ND

0.0 ND

0.3 ND

Cascades on Clyde

Jan 2017 May 2017

25.8 15.0

8.0 7.8

85.0 115.0

0.4 1.0

8.9 14.4

1.4 1.9

60.6 24.4

0.1 0.1

105 ND

0.7 ND

0.2 ND

ND = Not determined.

equilibrate, the overlying water was circulated in water tank. The following day, 15NO3− was added to the overlying water of cores of the wet treatment to measure DNF and DNRA while the cores of control were covered in the aerated site water until the end of the experiment. Rates of DNF and DNRA from wet treatment were measured when the DO was within ± 30 µmol-O2/L of the in situ oxygen level (Nielsen, 1992; Risgaard-Petersen et al., 1995; Dalsgaard et al., 2000). A 12 mL filtered (0.2 µm PES filter, Sartorius) NOx sample was collected to determine the initial NOx concentration after the addition of 15N-NO3 in the overlying water column. Water removed was replaced with site water and the cores were sealed. Before sacrificing each intact core, the DO was measured, and then 1 mL of 50% (w/v) ZnCl2 was added to the overlying water column to ensure microbial activity ceased. The core was gently slurried to minimize loss of 15N2 and allowed to settle for a minute before a N2 gas sample was collected in a 12 mL Exetainer (Labco) ensuring there were no bubbles in the sampling procedure or exetainer, which was then preserved with 250 µL of 50% (w/v) ZnCl2 until analysis. For DNRA, 20 mL of homogenized slurry was transferred into a 50 mL centrifuge tube (Falcon) and frozen until extraction with KCl and subsequent conversion of 15NH4+ to 15N2 via alkaline hypobromite (Risgaard-Petersen et al., 1995; Roberts et al., 2014). A random core was slurried at approximate time intervals of 1, 3, 5, and 7 h during which time the DO reduced by no more than 20% of the initial DO.

2000). The difference between the wet treatment and the control was the length of incubation period. Briefly, the sediments for the wet treatment were kept with an overlying water column in water tank for 2 days to determine the initial in situ nutrient fluxes and rates of DNF and DNRA. The sediments for the control were kept with an overlying water column in water tank for the duration of the experiment (9 days). This was to determine whether drying and rewetting cores over the same duration would lead to a change in nutrient fluxes and rates of DNF and DNRA. For the dry-rewet treatment, the water column was removed and sediments were kept drying for 7 days before rewetting for one day for measuring the nutrient fluxes and rates of DNF and DNRA together with the control (Fig. 1). The tank of site water was aerated using aerator to mimic the in situ O2 conditions (% DO saturation, Table 2) and the sediment cores were allowed to equilibrate overnight in dark conditions. The following day, fluxes of NH4+, FRP and NOx and sediment oxygen demand (SOD) were measured (Dalsgaard et al., 2000). A 12 mL filtered (0.2 µm PES filter, Sartorius) sample was collected to measure fluxes of NH4+, FRP and NOx (Fig. 1). Release of nutrients from sediments to the water column was indicated by a positive flux while a negative flux indicated uptake by the sediment. After completing the flux measurements, sediment cores for the wet treatment and control were covered in aerated site water overnight at in situ oxygen and temperature conditions to allow the sediment to re3

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In the dry-rewet treatment the overlying water from sediment cores was removed, care was taken to remove the overlying water to not disturb the sediment surface. The sediments were then dried under a halogen light bank with an average light intensity of ~450 µmol photons/m2/s to replicate the natural sunlight for 7 days. After 7 days, dried cores of the dry-rewet treatment were transferred into the aerated site water carefully to not disturb the sediment surface and left with cores of the control treatment and allowed to equilibrate overnight before experiments. The next day (day 8 of the experiment) nutrient fluxes and the following day (day 9 of the experiment) rates of DNF and DNRA were measured from the dry-rewet treatment and control as stated for the wet treatment. Rates of DNF and DNRA from intact cores were measured from the reduction of added 15NO3− as mentioned previously. The final concentrations of 15NO3− (15N atom%, 99.6%, Novachem Pty Ltd, Australia) added to the overlying water ranged from 50 to 100 µM depending on the volume of overlying water in each core. According to this technique and modified from Bernard et al. (2015), the rates at which 29N2 and 30N2 produced from 15NO3− are quantified and used to calculate 14N2 production rates (D14). The details of the calculation of DNF and DNRA in intact cores are in Rahman et al. (2019b). To eliminate the effect of concentrations of NO3− on NO3− reduction pathways and perform ANCOVA tests, potential rates of DNF (D15) and DNRA (DNRA15) in intact cores have been reported here.

and surface water by flow injection analysis (Lachat Quickchem 8000 Flow Injection Analyser with spectrophotometric detection) according to APHA (2005) methods. Matrix spikes and standard reference materials (SRM) were prepared and analyzed according to APHA (2005) and were found within the limits of 100 ± 10%. Denitrification was estimated by measuring the excess 15N2 accumulated in the headspace, while DNRA was estimated from the accumulation of 15NH4+ (Nielsen, 1992; Risgaard-Petersen et al., 1995; Dalsgaard et al., 2000) after conversion to 15N2 using the alkaline hypobromite method (Risgaard-Petersen et al., 1995; Roberts et al., 2014). Ammonium in samples was extracted from the slurry with 1:1 (v/v) 2 M KCl, shaken at 120 rpm for 1 h and centrifuged at 2000 rpm for 10 min. 8 mL of filtered extractant was transferred into a 12.5 mL Exetainer (Labco) and was purged with helium to eliminate background N2 followed by the addition of 200 µL of alkaline hypobromite to convert the 15NH4+ to 15N2. After adding hypobromite, samples were shaken at 130 rpm for 24 h prior to analysis to ensure all of the 15NH4+ was converted to 15N2 (Risgaard-Petersen et al., 1995). The hypobromite conversion of 15NH4+ provided a recovery of 100 ± 2–103 ± 0.1%. Isotope analysis of 15N2 was undertaken using a Sercon 20–22 continuous flow isotope ratio mass spectrometer (IRMS).

2.3. Slurry incubations for temperature experiments

Statistical analyses were conducted using SigmaPlot (SigmaPlot 13.0) and R (Version: R.3.2.0). Differences in potential rates of DNF and DNRA in slurry samples and their ratio from three treatments were determined using one-way analysis of variance (one-way ANOVA). In addition, the effect of three treatments on rates of DNF and DNRA in intact cores in drying-rewetting experiments were investigated through analysis of covariance (ANCOVA). ANCOVA was executed in R (Version: R.3.2.0). Error bars for rates reported in figures are ± 1 standard error of the slope.

2.5. Statistical analysis

The top 1 cm of the sediment core was extruded and mixed with site water to prepare 10% slurry samples (v/v); 8 mL slurry samples were transferred into 12.5 mL vials (Exetainer, Labco), sealed then purged with helium for 5 min to remove all oxygen. All Exetainers were preincubated overnight in the dark and shaken at 130 rpm to ensure that all the background 14NO3− and remaining oxygen were exhausted (Meyer et al., 2005; Roberts et al., 2014). Triplicate slurry samples were then amended with 15NO3− to a final concentration of 100 µM and incubated at four different temperatures: 7.5, 15, 25 and 35 °C. Samples for DNF and DNRA measurement were terminated by adding 250 μL of 50% ZnCl2 at 0, 10, 30, 60, 120 and 180 min after the addition of 15 NO3− (Meyer et al., 2005; Roberts et al., 2014). Rates of DNF and DNRA in slurries are expressed as ‘potential’ rates under optimal conditions of nitrate and carbon substrates. The activation energy for each process was calculated from the temperature dependence of the rates of DNF and DNRA. The slope of the plot of inverse temperature versus natural log of rates of DNF or DNRA provided an estimate of activation energy according to the Arrhenius equation (Rysgaard et al., 2004; Canion et al., 2014):

−Ea 1 ⎞ ⎛ ⎞ + ln(A) ln(k ) = ⎛ ⎝ R ⎠⎝T ⎠

3. Results 3.1. Surface water physico-chemical parameters Physico-chemical parameters were recorded in the surface water at all sites during each sampling occasion from December 2016 to May 2017 are provided in Table 2. The lowest autumn surface water temperature of 15 °C was observed in May 2016 at the Cascades on Clyde wetland while the highest summer surface water temperature of 28 °C was recorded in January 2017 at the Koolamara Blvd wetland. Surface water pH was nearly neutral to slightly basic ranging from 6.6 to 8.8. DO saturation during the study period ranged from 62 to 115%. Electrical conductivity of the surface water ranged from 0.2 to 1.1 mS cm−1 whereas turbidity ranged from 5.8 to 28.0 NTU. Concentrations of NH4+ in surface water ranged from 0.7 to 27 µM, NOx ranged from 0.3 to 114 µM and FRP ranged from 0.1 to 0.6 µM. However, in the sediment porewater, NH4+ ranged from 105 to 701 µM, NOx ranged from below the detection limit of 0.1 up to 0.7 µM and FRP ranged from 0.2 to 3.5 µM.

(1)

where k is the reaction rate, Ea is the activation energy (J mol−1), R is the gas constant (8.314 J mol−1 K−1), T is the absolute temperature (K) −Ea and A is the Arrhenius constant. R is equal to the slope of the plot of inverse temperature versus natural log of rates of DNF or DNRA. In addition, the temperature coefficient Q10, defined as the factor by which the reaction rate increases with a temperature increase of 10 °C was calculated according to the following equation (Rysgaard et al., 2004; Appelboom et al., 2006):

( )

3.2. Nutrient fluxes and sediment oxygen demand in intact cores Nutrient fluxes across the sediment-water interface and SOD were measured on samples collected between December 2016 and May 2017 (Fig. 2). Fluxes were measured on both day 1 and day 8 from the control and the dry-rewet treatments. NH4+ was released from the sediments on all occasions. The highest NH4+ efflux of 150 ± 20 µmol m−2 h−1 was observed from the sediments of Namatjira Reserve wetland before drying whereas the lowest NH4+ efflux of 23 ± 2 µmol m−2 h−1 was observed from the control of the Koolamara Blvd wetland after 8 days of incubations. Release of NH4+ from the sediments of the controls of all

10

Q10

k 2 (T 2 − T 1) =⎛ ⎞ ⎝ k1 ⎠

(2)

where k1 and k2 are the rates at temperatures T1 and T2, respectively. 2.4. Sample analysis Concentrations of NH4+, FRP and NOx were measured in porewater 4

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Fig. 2. Fluxes of NH4+ (a), NOx (b) and FRP (c) and SOD (d) between sediment and water column from intact core incubations of wet and dry-rewet treatments and control. D-1 for day 1 and D-8 for day 8. Error bars represent ± 1 SE. Positive values indicate fluxes out of the sediments while negative values indicate fluxes from water column into the sediments.

but the Namatjira Reserve wetland whereas release of NH4+ from the sediments of dry-rewet treatment of all wetlands decreased gradually with the length of the incubation. Unlike NH4+, FRP was taken up by the sediments in all treatments at all sampling dates with fluxes ranging from −2.0 ± 0.8 to −16 ± 3 µmol m−2 h−1. NOx was generally removed by sediments from the water column with the highest rate of −75 ± 9 µmol m−2 h−1 from the Namatjira Reserve wetland. However, in a few cases, NOx was released from the sediments into the water column after drying and rewetting the sediments, with the highest efflux rate of 42 ± 1 µmol m−2 h−1 from the Namatjira Reserve wetland. The SOD during the study period ranged from −71 ± 7 µmol O2 m−2 h−1 after rewetting the sediments of the Cascades on Clyde wetland to −450 ± 20 µmol O2 m−2 h−1 from the control of the Huntingdale Road wetland on day 1. The SOD from the Huntingdale Road and the Namatjira Reserve wetlands were higher compared to the Koolamara Blvd and the Cascades on Clyde wetlands. SOD decreased from the control sediments over time as well as from sediments after drying.

control of the Huntingdale Road wetland (Fig. 3b). Rates of DNRA in the dry-rewet treatment were significantly lower (p < 0.05) compared to the control from the Huntingdale Road and the Cascades on Clyde wetlands. However, rates of DNRA in the wet treatment were significantly lower (p < 0.05) compared to the control only from the Cascades on Clyde wetland. The DNF:DNRA ratio of higher than 1 indicates that DNF predominated in all three treatments except for the control of the Huntingdale Road wetland (Fig. 3c). The percentage contribution of DNRA to total nitrate reduction in intact cores ranged from 30 ± 15% in the wet treatment to 34 ± 24% in the control. 3.4. Denitrification and DNRA in slurries at different temperatures Potential rates of DNF in slurry samples ranged from 3.6 ± 0.1 µmol L slurry−1 h−1 at 7.5 °C to 160 ± 20 µmol L slurry−1 h−1 at 35 °C (Fig. 4a). All treatments showed approximately a doubling of reaction rate for every 10 °C increase in temperature (Table 3). The Huntingdale Road wetland showed the highest DNF rates at each temperature whereas the Koolamara Blvd wetland showed the lowest rates of DNF at each but 7.5 °C temperature. Activation energies (Ea, Table 3) for DNF at the Namatjira Reserve and Koolamara Blvd wetlands were quite similar of 41 kJ mol−1 whereas the Cascades on Clyde had the highest activation energy of 63.5 kJ mol−1. Potential rates of DNRA in slurries were always lower than DNF and ranged from 0.4 ± 0.3 µmol L slurry−1 h−1 at 7.5 °C to 28 ± 1 µmol L slurry−1 h−1 at 35 °C (Fig. 4b). Similar to DNF, all treatments showed approximately a doubling of reaction rate for every 10 °C increase in temperature (Table 3). Rates of DNRA from the Koolamara Blvd wetland were lower than the rates from the other three wetlands. The activation energies for DNRA were highly variable ranging from 50 to 107 kJ mol−1 with the lowest activation energy for the Namatjira Reserve and the highest for the Cascades on Clyde wetland.

3.3. Denitrification and DNRA in intact cores The rates of DNF (D15) measured in intact cores ranged from 5.0 ± 0.3 µmol N2 m−2 h−1 in the dry-rewet treatment of the Cascades on Clyde wetland to 90 ± 20 µmol-N2 m−2 h−1 in the wet treatment of the Huntingdale Road wetland (Fig. 3a). Rates of DNF in the dry-rewet treatment were significantly lower (p < 0.05) compared to the control from all but the Koolamara Blvd wetland. In contrast, rates of DNF in the wet treatment were significantly higher (p < 0.05) compared to the control only from the Cascades on Clyde wetland. Rates of DNRA (DNRA15) in intact cores were also variable ranging from 2.0 ± 0.2 µmol NH4+ m−2 h−1 in the dry-rewet treatment of the Cascades on Clyde wetland to 110 ± 40 µmol-NH4+ m−2 h−1 in the 5

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Fig. 3. Potential rates of denitrification (DNF, a), DNRA (b) and denitrification (DNF):DNRA ratio (c) from three treatments in intact core incubations. Error bars represent ± 1 SE.

Fig. 4. Potential rates of denitrification (a), DNRA (b) and denitrification:DNRA ratio (c) from slurry samples amended with 15N-NO3− and incubated at 7.5, 15, 25 and 35 °C. Error bars represent ± 1 SE (n = 3). Asterisk (*) denotes significant differences between 7.5 °C and other three temperatures.

The DNF:DNRA ratio of higher than 1 indicates that DNF dominated in slurries at all temperatures (Fig. 4c). The percent contribution of DNRA to total nitrate reduction in slurries ranged from 12 ± 2% at 7.5 °C to 24 ± 9% at 35 °C.

sediments on rates of denitrification (DNF) and DNRA. A decrease in DNF due to drying of sediments has also been observed in previous studies (Fromin et al., 2010; Sgouridis et al., 2011). For example, relatively higher rates of DNF were observed in the permanently flooded compared to the intermittent flooded zones in constructed riverine marsh wetlands in the Midwest USA (Hernandez and Mitsch, 2007). Reduced rates of DNF were observed from samples collected under drier conditions compared to those collected under water saturated

4. Discussion 4.1. Effect of drying-rewetting of sediments on denitrification and DNRA Our experiments suggest a clear effect of drying-rewetting of 6

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1994; Cui and Caldwell, 1997). Drying can result in the breakdown of the physical structure of sediments and hence may expose physically protected organic matter (Adu and Oades, 1978; Lundquist et al., 1999). This formerly inaccessible organic matter could then be quickly mineralized by the microbial community (Appel, 1998). Schimel et al. (1999) revealed that the microbial biomass and respiration on birch litter were altered by the frequency of the drying-rewetting events. The drying phase killed sediment bacteria and thus enhanced respiration by making more organic matter available to the living organisms during the rewetting phase (Scholz et al., 2002). Although rates of respiration increased as indicated by increased NH4+ fluxes (Fig. 2a), rates of both DNF and DNRA decreased significantly (p < 0.05) from the dry-rewet treatment compared to the control possibly due to the oxic condition of sediments. Similar to our results, an increase in mineralization was also observed due to drainage following aeration of wet soils (Cabrera, 1993; Updegraff et al., 1995; Bridgham et al., 1998), however, rates of DNF dramatically decreased because of the decrease in anoxic regions through drying (Groffman and Tiedje, 1988; Seitzinger, 1994). Moreover, stimulation of DNF due to drying-rewetting caused primarily by decreased levels of O2 in soils due to aerobic respiration rather than by the availability of more substrates for the denitrifiers (Groffman and Tiedje, 1988). Therefore, we suggest that lower rates of DNF and DNRA from the dry-rewet treatment compared to the wet treatment were mainly due to the oxic condition of sediments. In 50% of cases, the DNF:DNRA ratio from the dry-rewet treatment were higher compared to the control. This indicates that DNRA significantly responds to drying (Fig. 3c). This is because obligately anaerobic fermentative bacteria that carry out DNRA are more susceptible to small changes in oxygen and hence DNRA enzyme activity was more inhibited by drying than DNF (Smith and Parsons, 1985; Groffman and Tiedje, 1988; Burgin and Hamilton, 2007; KoopJakobsen and Giblin, 2010). This suggests that the oxic condition of sediments resulted in a change in the microbial community structure through death of nitrate reducing bacteria and in particular DNRA bacteria during drying period (Fierer and Schimel, 2002; Fierer et al., 2003; Fay et al., 2016).

Table 3 Activation energy (Ea) and temperature coefficient (Q10) for all four wetlands. Values of Q10 were calculated between 15 and 25; and 25 and 35 °C. Huntingdale

Namatjira

Koolamara

Cascades

DNF

Ea (kJ mol−1) Q10 (15–25 °C) Q10 (25–35 °C)

53 ± 5 2.2 1.9

41 ± 7 1.3 1.5

41.0 ± 0.7 1.7 1.3

64 ± 4 2.4 1.5

DNRA

Ea (KJ mol−1) Q10 (15–25 °C) Q10 (25–35 °C)

62 ± 6 2.3 2.1

50 ± 8 1.4 2.5

57 ± 9 1.6 1.2

110 ± 10 3.4 1.7

conditions in a temperate re-connected floodplain. The differences in rates of DNF were mainly caused by the differences in moisture content and the resulting fluctuating O2 conditions (Sgouridis et al., 2011). In natural riverine wetlands in Minnesota, DNF potential during summer was higher in zones where the soil surface was submerged compared to zones slightly raised above the water table (Johnston et al., 2001). In tidal salt marshes in South England, consistently higher rates of DNF were observed from low marsh compared to high marsh zones and mudflats (Koch et al., 1992). Flooded surface horizons showed higher average rates of DNF than terrestrial soils in the Virginia Coastal Plain (Pavel et al., 1996). Anoxic conditions in permanently flooded sediments in these wetlands were the main controlling factor of the rates of DNF. In permanently flooded sediments, DNF is enhanced by the supply of organic matter as well as the creation of aerobic zones for nitrification by the root zones of emergent macrophyte vegetation and eventually supplying nitrate to the anoxic zone for DNF (Hernandez and Mitsch, 2007). To our knowledge, there is no study on the effect of drying-rewetting of sediments on DNRA to compare our results. In this regard, this study provides a unique insight into the effect of drying-rewetting on DNRA and hence N cycling in constructed urban wetlands. We hypothesised that drying of sediments resulted in death of obligately anaerobic bacteria due to switching from anoxic to oxic conditions and hence a change in the microbial community; primarily DNRA bacteria. It has previously been shown that denitrifier communities are stable over short-term water table fluctuations in wetlands because of their communal versatility (Fromin et al., 2010; Song et al., 2010; McKew et al., 2011). Drying-rewetting conditions of sediments can change the response of microbial processes such as DNF and DNRA not only due to changes in the availability of nutrients but also because changes in microbial communities due to physiological stress (Fierer and Schimel, 2002; Fierer et al., 2003; Fromin et al., 2010). For example, fast growing microorganisms have been identified to be more vulnerable to drying-rewetting stress compared to slow growing microorganisms, presumably because of contrasts in characteristics of the cell wall (Bottner, 1985; Van Gestel et al., 1993a,b). Rapid changes in soil water potential may favour gram-positive bacteria and fungi, which have thicker, more inflexible cell walls and compatible solutes that improve osmoregulatory abilities (Schimel et al., 1989; Kempf and Bremer, 1998; Schimel et al., 1999). However, other studies have reported that repeated drying-rewetting may favour fast growing organisms that are fit for quick development on the labile substrates discharged into the soil through a rewetting event (Jager and Bruins, 1975; Lund and Goksøyr, 1980; Scheu and Parkinson, 1994; Denef et al., 2001). Moreover, rewetting of the dried sediments would result in increased mineralization of sediment organic matter with increased respiration rates. Higher fluxes of NH4+ and after 8 days, similar SOD from the dry-rewet treatment compared to the control sediments also supports this hypothesis of increased respiration in dried and rewetted sediments. It has previously been shown that rewetting of a dry soil resulted in an increase in C and N mineralization rates compared to an unstressed control (Birch, 1958; Sørensen, 1974; Scheu and Parkinson,

4.2. Effect of temperature on denitrification and DNRA Our results from slurries incubated at four different temperatures revealed that temperature had a positive effect on both DNF and DNRA. In agreement with our results, rates of DNF in intact cores increased with increasing temperatures in previous studies (Seitzinger, 1988; Amatya et al., 2009; Veraart et al., 2011). Rates of DNRA in two constructed freshwater lakes in Poland were positively correlated with temperature (Tomaszek and Rokosz, 2007; Gruca-Rokosz et al., 2009). In this study, the relative proportion of DNF and DNRA at different temperatures illustrates that the relative increase in DNRA was more than DNF with increasing temperatures, which is in agreement with previous studies on estuarine and marine ecosystems (King and Nedwell, 1984; Ogilvie et al., 1997; Dong et al., 2011). Although several studies have described the mechanisms for higher rates of DNF and DNRA with rising temperature (Cavari and Phelps, 1977; Bachand and Horne, 1999; Veraart et al., 2011), the effect of temperature on DNF and DNRA seems to vary substantially between systems. For example, Brin et al. (2015) observed that rates of both DNF and DNRA decreased in warmed treatments, potentially reflecting carbon loss due to increased respiration with warming. In comparison to DNF, most studies have found that DNRA potential increased with increasing temperatures (King and Nedwell, 1984; Ogilvie et al., 1997; Tomaszek and Rokosz, 2007; Gruca-Rokosz et al., 2009; Dodsworth et al., 2011). In this study, our results also suggest that the increase in DNRA was more than DNF with increasing temperatures and this is because nitrate ammonifying bacteria are favoured by higher temperatures. The activation energy, Ea, of DNF (41–64 kJ mol−1) observed for four wetlands in this study (Table 3) is comparable to the range 7

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(60–98 kJ mol−1) reported for nitrate-reducing communities in salt marsh sediment (King and Nedwell, 1984), temperate permeable sediments (52–65 kJ mol−1) (Canion et al., 2014) and Arctic sediments (58–60 kJ mol−1) (Rysgaard et al., 2004; Canion et al., 2014). However, the Ea of DNF in this study is few times higher compared to the Ea of groundwater DNF beneath wetland soils (13 kJ mol−1) (Ribas et al., 2015). Furthermore, the Q10 values of DNF (1.3–2.4) in this study (Table 3) are also within the range of 1.3–3.0 observed by most studies for DNF although the Q10 coefficient is generally approximated as 2 indicating a 2-fold increase in rate for each 10 °C rise in temperature (Ellis et al., 1998; Martin et al., 2001; Lewis, 2002; Yu et al., 2006). The lower Ea values of DNF (41–64 kJ mol−1) in this study similar to the lower Ea values (52–65 kJ mol−1) from the polar and temperate sediments (Rysgaard et al., 2004; Canion et al., 2014) indicate the significant contribution of psychrophilic to psychrotolerant populations to DNF (Canion et al., 2014). Compared to DNF, there is a paucity of Ea values of DNRA in the literatures with no values reported on wetlands. The values of Ea of DNRA (50–107 kJ mol−1) observed for four wetlands in this study (Table 3) are 2–4 times higher compared to the Ea value of DNRA (kJ mol−1) observed from coastal, sandy tidal flat sediments (Kraft et al., 2014). However, the Q10 values of DNRA (1.2–3.4) in this study are mostly comparable to the Q10 value of DNRA (1.5) from coastal, sandy tidal flat sediments (Kraft et al., 2014). A wide range of Ea values for DNRA in this study is an indicative of the presence of both the psychrophilic and the mesophilic DNRA bacterial community (Canion et al., 2014). Sediments of the Cascades on Clyde wetland had the Ea value of 107 kJ mol−1, which is approximately twice compared to the other three wetlands and indicate the presence of an exclusively mesophilic community (Canion et al., 2014). King and Nedwell (1984) and JØrgensen (1989), observed that DNRA bacteria dominated over denitrifying bacteria in higher temperatures during summer whereas denitrifying bacteria dominated in cold autumn and winter. Ogilvie et al. (1997) demonstrated that only nitrate ammonifiers (Klebsiella and Enterobacter) were identified from estuarine sediments at 20 °C, whereas just denitrifiers below 20 °C, and the reason was the higher affinity of nitrate ammonifiers for nitrate than denitrifiers at higher temperatures. Moreover, Kelly-Gerreyn et al. (2001) observed from a model output that denitrifiers dominated at temperatures between 14 and 17 °C while nitrate ammonifiers dominated below 14 and above 17 °C. King and Nedwell (1984) observed that Pseudomonas sp., which reduce nitrate mainly to gaseous end product (N2 and N2O) and to some extent to nitrite (Macfarlane and Herbert, 1982; Herbert et al., 2011) dominated at 10 °C whereas metabolically fermentative bacteria such as Vibrio sp. dominated at 25 °C which convert nitrate to nitrite and ammonium (Cole and Brown, 1980).

urban heat island effect may negatively affect the nitrogen processing capacity of the urban stormwater wetlands. 6. Conclusion Results from drying-rewetting experiments showed that rates of both DNF and DNRA decreased significantly due to drying of sediments. However, relatively higher DNF:DNRA from the dry-rewet treatment compared to the control suggests that drying condition of sediments resulted in a reduction in DNRA more than DNF. We suggest that drying the sediments resulted in a change in microbial community structure through bacterial death, which eventually caused a decrease in DNF and DNRA. Moreover, a higher inhibition of DNRA suggests that nitrate ammonifying bacteria were affected more by drying conditions than denitrifiers. In addition to drying, the explicit investigation of the effect of temperature on the partitioning between DNF and DNRA in this study revealed that rates of both DNF and DNRA increased with increasing temperatures. Although DNF predominated at all temperatures, a 2–62% increase in DNRA compared to a 2–12% increase in DNF for an increase of 8–28 °C in temperature suggests that the increase in DNRA was more than DNF. We infer that higher temperature favours nitrate ammonifying bacteria and hence rates of DNRA increase more rapidly compared to DNF with increasing temperature. Declaration of Competing Interest The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Acknowledgments We thank Lee James, the Supervisor of Natural Resource Areas, Parks Department (City of Kingston) and his team for providing access to, and permission for, sampling the Namatjira Reserve and John Erwin, the Bushland Management Officer (Knox City Council), for his invaluable discussion on the Koolamara Blvd wetland. Bipasa Akter is thanked for her assistance with fieldwork. Md Moklesur Rahman is grateful to the Co-operative Research Council for the Water Sensitive Cities (CRCWSC) Program B2.2-2.3 for funding support in carrying out this project and Melbourne Water for providing important information on the studied wetlands. Appendix A. Supplementary data Supplementary data to this article can be found online at https:// doi.org/10.1016/j.ecoleng.2019.105586.

5. Implications of this study References This study revealed that the drying-rewetting events of sediments have a significant effect on the nitrogen processing pathways in the constructed urban stormwater wetlands. Drying-rewetting condition of sediments affected both removal and recycling of nitrogen in all wetlands studied. Although rates of DNF and DNRA were measured only once for a short period for the drying-rewetting experiments mainly due to the time constraints of this project, the ANCOVA results suggested that in 75% cases, DNF and in 50% cases, DNRA decreased significantly (p < 0.05) from the dry-rewet treatment compared to the control. Therefore, we suggest that drying-rewetting cycles will lead to reduced total rates of nitrogen removal in water treatment wetlands. Results from the temperature experiments suggested that both removal and recycling of nitrogen were enhanced by increasing temperature in slurries. Typically, a decrease in the DNF:DNRA with increasing temperature was observed, suggesting that recycling of nitrogen was stimulated more than removal in the wetlands studied. This implies that higher temperatures in summer due to seasonal variabilities and the

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